TOXICOLOGICAL PROFILE FOR POLYBROMINATED DIPHENYL ETHERS (PBDEs) U.S. DEPARTMENT OF HEALTH AND HUMAN SERVICES Public Health Service Agency for Toxic Substances and Disease Registry March 2017 iiPBDEsDISCLAIMER Use of trade names is for identification only and does not imply endorsement by the Agency for Toxic Substances and Disease Registry, the Public Health Service, or the U.S. Department of Health and Human Services. UPDATE STATEMENT A Toxicological Profile for Polybrominated Diphenyl Ethers, Draft for Public Comment was released in September 2015. This edition supersedes any previously released draft or final profile. Toxicological profiles are revised and republished as necessary. For information regarding the update status of previously released profiles, contact ATSDR at: Agency for Toxic Substances and Disease Registry Division of Toxicology and Human Health Sciences Environmental Toxicology Branch 1600 Clifton Road NE Mailstop F-57 Atlanta, Georgia 30329-4027 iiiPBDEs ivPBDEsThis page is intentionally blank. FOREWORD This toxicological profile is prepared in accordance with guidelines* developed by the Agency for Toxic Substances and Disease Registry (ATSDR) and the Environmental Protection Agency (EPA). The original guidelines were published in the Federal Register on April 17, 1987. Each profile will be revised and republished as necessary. The ATSDR toxicological profile succinctly characterizes the toxicologic and adverse health effects information for these toxic substances described therein. Each peer-reviewed profile identifies and reviews the key literature that describes a substance's toxicologic properties. Other pertinent literature is also presented, but is described in less detail than the key studies. The profile is not intended to be an exhaustive document; however, more comprehensive sources of specialty information are referenced. The focus of the profiles is on health and toxicologic information; therefore, each toxicological profile begins with a public health statement that describes, in nontechnical language, a substance's relevant toxicological properties. Following the public health statement is information concerning levels of significant human exposure and, where known, significant health effects. The adequacy of information to determine a substance's health effects is described in a health effects summary. Data needs that are of significance to protection of public health are identified by ATSDR. Each profile includes the following: (A) The examination, summary, and interpretation of available toxicologic information and epidemiologic evaluations on a toxic substance to ascertain the levels of significant human exposure for the substance and the associated acute, subacute, and chronic health effects; (B) A determination of whether adequate information on the health effects of each substance is available or in the process of development to determine levels of exposure that present a significant risk to human health of acute, subacute, and chronic health effects; and (C) Where appropriate, identification of toxicologic testing needed to identify the types or levels of exposure that may present significant risk of adverse health effects in humans. The principal audiences for the toxicological profiles are health professionals at the Federal, State, and local levels; interested private sector organizations and groups; and members of the public. This profile reflects ATSDR’s assessment of all relevant toxicologic testing and information that has been peer-reviewed. Staffs of the Centers for Disease Control and Prevention and other Federal scientists have also reviewed the profile. In addition, this profile has been peer-reviewed by a nongovernmental panel and was made available for public review. Final responsibility for the contents and views expressed in this toxicological profile resides with ATSDR. Patrick N. Breysse, Ph.D., CIH Director, National Center for Environmental Health and Agency for Toxic Substances and Disease Registry Centers for Disease Control and Prevention vPBDEs *Legislative Background The toxicological profiles are developed under the Comprehensive Environmental Response, Compensation, and Liability Act of 1980, as amended (CERCLA or Superfund). CERCLA section 104(i)(1) directs the Administrator of ATSDR to “…effectuate and implement the health related authorities” of the statute. This includes the preparation of toxicological profiles for hazardous substances most commonly found at facilities on the CERCLA National Priorities List and that pose the most significant potential threat to human health, as determined by ATSDR and the EPA. Section 104(i)(3) of CERCLA, as amended, directs the Administrator of ATSDR to prepare a toxicological profile for each substance on the list. In addition, ATSDR has the authority to prepare toxicological profiles for substances not found at sites on the National Priorities List, in an effort to “…establish and maintain inventory of literature, research, and studies on the health effects of toxic substances” under CERCLA Section 104(i)(1)(B), to respond to requests for consultation under section 104(i)(4), and as otherwise necessary to support the site-specific response actions conducted by ATSDR. viPBDEs PBDEs viiQUICK REFERENCE FOR HEALTH CARE PROVIDERSToxicological Profiles are a unique compilation of toxicological information on a given hazardoussubstance. Each profile reflects a comprehensive and extensive evaluation, summary, and interpretationof available toxicologic and epidemiologic information on a substance. Health care providers treating patients potentially exposed to hazardous substances will find the following information helpful for fastanswers to often-asked questions.Primary Chapters/Sections of InterestChapter 1: Public Health Statement: The Public Health Statement can be a useful tool for educatingpatients about possible exposure to a hazardous substance. It explains a substance’s relevanttoxicologic properties in a nontechnical, question-and-answer format, and it includes a review ofthe general health effects observed following exposure.Chapter 2: Relevance to Public Health: The Relevance to Public Health Section evaluates, interprets,and assesses the significance of toxicity data to human health.Chapter 3: Health Effects: Specific health effects of a given hazardous compound are reported by typeof health effect (death, systemic, immunologic, reproductive), by route of exposure, and by length of exposure (acute, intermediate, and chronic). In addition, both human and animal studies arereported in this section.NOTE: Not all health effects reported in this section are necessarily observed in the clinicalsetting. Please refer to the Public Health Statement to identify general health effects observedfollowing exposure.Pediatrics: Four new sections have been added to each Toxicological Profile to address child health issues:Chapter 1 How Can (Chemical X) Affect Children?Chapter 1 How Can Families Reduce the Risk of Exposure to (Chemical X)?Section 3.7 Children’s SusceptibilitySection 6.6 Exposures of ChildrenOther Sections of Interest:Section 3.8 Biomarkers of Exposure and EffectSection 3.11 Methods for Reducing Toxic EffectsATSDR Information CenterPhone: 1-800-CDC-INFO (800-232-4636) or 1-888-232-6348 (TTY)Internet: http://www.atsdr.cdc.govThe following additional material is available online at www.atsdr.cdc.gov:Case Studies in Environmental Medicine—Case Studies are self-instructional publications designed to increase primary care provider’s knowledge of a hazardous substance in the environment and toaid in the evaluation of potentially exposed patients. Managing Hazardous Materials Incidents is a three-volume set of recommendations for on-scene (prehospital) and hospital medical management of patients exposed during a hazardous materials PBDEs viiiincident. Volumes I and II are planning guides to assist first responders and hospital emergencydepartment personnel in planning for incidents that involve hazardous materials. Volume III— Medical Management Guidelines for Acute Chemical Exposures—is a guide for health care professionals treating patients exposed to hazardous materials.Fact Sheets (ToxFAQs™) provide answers to frequently asked questions about toxic substances.Other Agencies and OrganizationsThe National Center for Environmental Health (NCEH) focuses on preventing or controlling disease, injury, and disability related to the interactions between people and their environment outside theworkplace. Contact: NCEH, Mailstop F-29, 4770 Buford Highway, NE, Atlanta,GA 30341-3724 • Phone: 770-488-7000 • FAX: 770-488-7015.The National Institute for Occupational Safety and Health (NIOSH) conducts research on occupationaldiseases and injuries, responds to requests for assistance by investigating problems of health andsafety in the workplace, recommends standards to the Occupational Safety and Health Administration (OSHA) and the Mine Safety and Health Administration (MSHA), and trainsprofessionals in occupational safety and health. Contact: NIOSH, 395 E Street, S.W., Suite 9200, Patriots Plaza Building, Washington, DC 20201 • Phone: (202) 245-0625 or 1-800-CDC-INFO (800-232-4636).The National Institute of Environmental Health Sciences (NIEHS) is the principal federal agency forbiomedical research on the effects of chemical, physical, and biologic environmental agents on human health and well-being. Contact: NIEHS, PO Box 12233, 104 T.W. Alexander Drive, Research Triangle Park, NC 27709 • Phone: 919-541-3212.Publically Available InformationThe Association of Occupational and Environmental Clinics (AOEC) has developed a network of clinics in the United States to provide expertise in occupational and environmental issues. Contact:AOEC, 1010 Vermont Avenue, NW, #513, Washington, DC 20005 • Phone: 202-347-4976 • FAX: 202-347-4950 • e-mail: AOEC@AOEC.ORG • Web Page: http://www.aoec.org/.The American College of Occupational and Environmental Medicine (ACOEM) is an association ofphysicians and other health care providers specializing in the field of occupational and environmental medicine. Contact: ACOEM, 25 Northwest Point Boulevard, Suite 700, ElkGrove Village, IL 60007-1030 • Phone: 847-818-1800 • FAX: 847-818-9266.The American College of Medical Toxicology (ACMT) is a nonprofit association of physicians withrecognized expertise in medical toxicology. Contact: ACMT, 10645 North Tatum Boulevard, Suite 200-111, Phoenix AZ 85028 • Phone: 844-226-8333 • FAX: 844-226-8333 • Web Page:http://www.acmt.net.The Pediatric Environmental Health Specialty Units (PEHSUs) is an interconnected system of specialists who respond to questions from public health professionals, clinicians, policy makers, and thepublic about the impact of environmental factors on the health of children and reproductive-agedadults. Contact information for regional centers can be found at http://pehsu.net/findhelp.html. ixPBDEsThe American Association of Poison Control Centers (AAPCC) provide support on the prevention and treatment of poison exposures. Contact: AAPCC, 515 King Street, Suite 510, Alexandria VA22314 • Phone: 701-894-1858 • Poison Help Line: 1-800-222-1222. xPBDEsThis page is intentionally blank. xiPBDEsCONTRIBUTORSCHEMICAL MANAGER(S)/AUTHOR(S):Hana R. Pohl, M.D., Ph.D.ATSDR, Division of Toxicology and Human Health Sciences, Atlanta, GAMarc Odin, M.S., DABTPeter McClure, Ph.D., DABTKimberly Zaccaria, Ph.D., DABTFernando Llados, Ph.D.Mary Kawa, M.A.Mario Citra, Ph.D.SRC, Inc., North Syracuse, NYTHE PROFILE HAS UNDERGONE THE FOLLOWING ATSDR INTERNAL REVIEWS:1. Health Effects Review. The Health Effects Review Committee examines the health effects chapter of each profile for consistency and accuracy in interpreting health effects and classifyingend points.2. Minimal Risk Level Review. The Minimal Risk Level Workgroup considers issues relevant tosubstance-specific Minimal Risk Levels (MRLs), reviews the health effects database of eachprofile, and makes recommendations for derivation of MRLs.3. Data Needs Review. The Environmental Toxicology Branch reviews data needs sections toassure consistency across profiles and adherence to instructions in the Guidance.4. Green Border Review. Green Border review assures the consistency with ATSDR policy. PBDEs xiiThis page is intentionally blank. PBDEs xiiiPEER REVIEWA peer review panel was assembled for polybrominated diphenyl ethers. The panel consisted of the following members:Dr. Stuart Harrad, Division of Environmental Health and Risk Management, School ofGeography, Earth, and Environmental Sciences, University of Birmingham, Edgbaston, Birmingham, United Kingdom;2. Dr. James R. Olson, Jr., Department of Chemistry, University at Buffalo, State University of NewYork, Buffalo, New York; and3. Dr. Christopher Metcalfe, Metcalfe C. Environmental and Resource Studies, Trent University, 1600 West Bank Drive, Peterborough, Ontario, Canada.These experts collectively have knowledge of polybrominated diphenyl ether’s physical and chemicalproperties, toxicokinetics, key health end points, mechanisms of action, human and animal exposure, and quantification of risk to humans. All reviewers were selected in conformity with the conditions for peerreview specified in Section 104(I)(13) of the Comprehensive Environmental Response, Compensation, and Liability Act, as amended.Scientists from the Agency for Toxic Substances and Disease Registry (ATSDR) have reviewed the peerreviewers' comments and determined which comments will be included in the profile. A listing of thepeer reviewers' comments not incorporated in the profile, with a brief explanation of the rationale for theirexclusion, exists as part of the administrative record for this compound. The citation of the peer review panel should not be understood to imply its approval of the profile's final content. The responsibility for the content of this profile lies with the ATSDR.1. PBDEs xivThis page is intentionally blank. CONTENTS DISCLAIMER .............................................................................................................................................. ii UPDATE STATEMENT ............................................................................................................................. iii FOREWORD ................................................................................................................................................ v QUICK REFERENCE FOR HEALTH CARE PROVIDERS .................................................................... vii CONTRIBUTORS ....................................................................................................................................... xi PEER REVIEW ......................................................................................................................................... xiii CONTENTS ................................................................................................................................................ xv LIST OF FIGURES ................................................................................................................................... xix LIST OF TABLES ..................................................................................................................................... xxi 1. PUBLIC HEALTH STATEMENT FOR POLYBROMINATED DIPHENYL ETHERS (PBDEs) ....... 1 2. RELEVANCE TO PUBLIC HEALTH ................................................................................................... 9 2.1 BACKGROUND AND ENVIRONMENTAL EXPOSURES TO PBDEs IN THE UNITED STATES ......................................................................................................................................... 9 2.2 SUMMARY OF HEALTH EFFECTS ......................................................................................... 11 2.3 MINIMAL RISK LEVELS (MRLs) ............................................................................................ 26 3. HEALTH EFFECTS .............................................................................................................................. 43 3.1 INTRODUCTION ........................................................................................................................ 43 3.2 DISCUSSION OF HEALTH EFFECTS BY ROUTE OF EXPOSURE ..................................... 44 3.2.1 Inhalation Exposure .............................................................................................................. 45 3.2.1.1 Death .............................................................................................................................. 45 3.2.1.2 Systemic Effects ............................................................................................................. 46 3.2.1.3 Immunological and Lymphoreticular Effects ................................................................ 55 3.2.1.4 Neurological Effects ...................................................................................................... 56 3.2.1.5 Reproductive Effects ...................................................................................................... 56 3.2.1.6 Developmental Effects ................................................................................................... 56 3.2.1.7 Cancer ............................................................................................................................ 57 3.2.2 Oral Exposure ........................................................................................................................ 57 3.2.2.1 Death .............................................................................................................................. 57 3.2.2.2 Systemic Effects........................................................................................................... 135 3.2.2.3 Immunological and Lymphoreticular Effects .............................................................. 175 3.2.2.4 Neurological Effects .................................................................................................... 181 3.2.2.5 Reproductive Effects .................................................................................................... 185 3.2.2.6 Developmental Effects ................................................................................................. 197 3.2.2.7 Cancer .......................................................................................................................... 236 3.2.3 Dermal Exposure ................................................................................................................. 240 3.2.3.1 Death ............................................................................................................................ 240 3.2.3.2 Systemic Effects........................................................................................................... 240 3.2.3.3 Immunological and Lymphoreticular Effects .............................................................. 243 3.2.3.4 Neurological Effects .................................................................................................... 243 3.2.3.5 Reproductive Effects .................................................................................................... 243 3.2.3.6 Developmental Effects ................................................................................................. 243 3.2.3.7 Cancer .......................................................................................................................... 243 3.3 GENOTOXICITY ...................................................................................................................... 243 3.4 TOXICOKINETICS ................................................................................................................... 246 3.4.1 Absorption ........................................................................................................................... 246 3.4.1.1 Inhalation Exposure ..................................................................................................... 246 PBDEs xv 3.4.1.2 Oral Exposure .............................................................................................................. 246 3.4.1.3 Dermal Exposure ......................................................................................................... 252 3.4.2 Distribution ......................................................................................................................... 253 3.4.2.1 Inhalation Exposure ..................................................................................................... 253 3.4.2.2 Oral Exposure .............................................................................................................. 253 3.4.2.3 Dermal Exposure ......................................................................................................... 260 3.4.3 Metabolism .......................................................................................................................... 260 3.4.4 Elimination and Excretion ................................................................................................... 267 3.4.4.1 Inhalation Exposure ..................................................................................................... 267 3.4.4.2 Oral Exposure .............................................................................................................. 268 3.4.4.3 Dermal Exposure ......................................................................................................... 270 3.4.5 Physiologically Based Pharmacokinetic (PBPK)/Pharmacodynamic (PD) Models ........... 270 3.5 MECHANISMS OF ACTION ................................................................................................... 273 3.5.1 Pharmacokinetic Mechanisms ............................................................................................. 273 3.5.2 Mechanisms of Toxicity ...................................................................................................... 276 3.5.3 Animal-to-Human Extrapolations ....................................................................................... 296 3.6 TOXICITIES MEDIATED THROUGH THE NEUROENDOCRINE AXIS ........................... 297 3.7 CHILDREN’S SUSCEPTIBILITY ............................................................................................ 299 3.8 BIOMARKERS OF EXPOSURE AND EFFECT ..................................................................... 308 3.8.1 Biomarkers Used to Identify or Quantify Exposure to PBDEs ........................................... 309 3.8.2 Biomarkers Used to Characterize Effects Caused by PBDEs ............................................. 310 3.9 INTERACTIONS WITH OTHER CHEMICALS ..................................................................... 311 3.10 POPULATIONS THAT ARE UNUSUALLY SUSCEPTIBLE ................................................ 314 3.11 METHODS FOR REDUCING TOXIC EFFECTS .................................................................... 316 3.11.1 Reducing Peak Absorption Following Exposure ............................................................. 316 3.11.2 Reducing Body Burden ................................................................................................... 317 3.11.3 Interfering with the Mechanism of Action for Toxic Effects .......................................... 318 3.12 ADEQUACY OF THE DATABASE ........................................................................................ 320 3.12.1 Existing Information on Health Effects of PBDEs .......................................................... 320 3.12.2 Identification of Data Needs ............................................................................................ 322 3.12.3 Ongoing Studies .............................................................................................................. 336 4. CHEMICAL AND PHYSICAL INFORMATION .............................................................................. 343 4.1 CHEMICAL IDENTITY ............................................................................................................ 343 4.2 PHYSICAL AND CHEMICAL PROPERTIES ......................................................................... 352 5. PRODUCTION, IMPORT/EXPORT, USE, AND DISPOSAL .......................................................... 359 5.1 PRODUCTION .......................................................................................................................... 359 5.2 IMPORT/EXPORT .................................................................................................................... 361 5.3 USE ............................................................................................................................................ 361 5.4 DISPOSAL ................................................................................................................................. 362 6. POTENTIAL FOR HUMAN EXPOSURE ......................................................................................... 365 6.1 OVERVIEW ............................................................................................................................... 365 6.2 RELEASES TO THE ENVIRONMENT ................................................................................... 366 6.2.1 Air ....................................................................................................................................... 366 6.2.2 Water ................................................................................................................................... 370 6.2.3 Soil ...................................................................................................................................... 372 6.3 ENVIRONMENTAL FATE ...................................................................................................... 373 6.3.1 Transport and Partitioning ................................................................................................... 373 6.3.2 Transformation and Degradation ........................................................................................ 376 PBDEs xvi 6.3.2.1 Air ................................................................................................................................ 376 6.3.2.2 Water ............................................................................................................................ 377 6.3.2.3 Sediment and Soil ........................................................................................................ 381 6.3.2.4 Other Media ................................................................................................................. 381 6.4 LEVELS MONITORED OR ESTIMATED IN THE ENVIRONMENT .................................. 381 6.4.1 Air ....................................................................................................................................... 382 6.4.2 Water ................................................................................................................................... 385 6.4.3 Sediment and Soil ............................................................................................................... 386 6.4.4 Other Environmental Media ................................................................................................ 390 6.5 GENERAL POPULATION AND OCCUPATIONAL EXPOSURE ........................................ 433 6.6 EXPOSURES OF CHILDREN .................................................................................................. 437 6.7 POPULATIONS WITH POTENTIALLY HIGH EXPOSURES .............................................. 439 6.8 ADEQUACY OF THE DATABASE ........................................................................................ 439 6.8.1 Identification of Data Needs ............................................................................................... 440 6.8.2 Ongoing Studies .................................................................................................................. 443 7. ANALYTICAL METHODS................................................................................................................ 445 7.1 BIOLOGICAL MATERIALS .................................................................................................... 447 7.2 ENVIRONMENTAL SAMPLES .............................................................................................. 450 7.3 ADEQUACY OF THE DATABASE ........................................................................................ 455 7.3.1 Identification of Data Needs ............................................................................................... 455 7.3.2 Ongoing Studies .................................................................................................................. 456 8. REGULATIONS, ADVISORIES, AND GUIDELINES ..................................................................... 457 9. REFERENCES .................................................................................................................................... 463 10. GLOSSARY ...................................................................................................................................... 527 APPENDICES A. ATSDR MINIMAL RISK LEVELS AND WORKSHEETS ............................................................. A-1 B. USER’S GUIDE .................................................................................................................................. B-1 C. ACRONYMS, ABBREVIATIONS, AND SYMBOLS ...................................................................... C-1 PBDEs xvii PBDEs xviiiThis page is intentionally blank. LIST OF FIGURES 3-1. Levels of Significant Exposure to Lower Brominated Diphenyl Ethers – Inhalation ........................ 49 3-2. Levels of Significant Exposure to Lower Brominated Diphenyl Ethers – Oral ............................... 103 3-3. Levels of Significant Exposure to Decabromodiphenyl Ether – Oral .............................................. 130 3-4. Structures and General Metabolic Scheme for Hydroxylated Metabolites of BDE 47 Produced by Human Liver Microsomes ........................................................................................................... 263 3-5. Structures and General Metabolic Scheme for Hydroxylated Metabolites of BDE 99 Produced by Human Liver Microsomes ........................................................................................................... 264 3-6. Structures and General Metabolic Scheme for Hydroxylated Metabolites of BDE 100 Produced by Human Liver Microsomes and Human CYP2B6 ........................................................................ 265 3-7. Conceptual Representation of a Physiologically Based Pharmacokinetic (PBPK) Model for a Hypothetical Chemical Substance .................................................................................................... 272 3-8. Existing Information on Health Effects of Polybrominated Diphenyl Ethers (PBDEs) ................... 321 6-1. Total On- and Off-Site Releases of Decabromodiphenyl Ether, 1998–2012 ................................... 369 6-2. Temporal Trends in the Concentrations (ng/g Lipid Weight) of PBDE Congeners in Lake Trout from Lake Ontario, Canada .............................................................................................................. 403 PBDEs xix PBDEs xxThis page is intentionally blank. LIST OF TABLES 3-1. Levels of Significant Exposure to Lower Brominated Diphenyl Ethers – Inhalation ........................ 46 3-2. Levels of Significant Exposure to Lower Brominated Diphenyl Ethers – Oral ................................. 58 3-3. Levels of Significant Exposure to Decabromodiphenyl Ether – Oral .............................................. 110 3-4. Levels of Significant Exposure to Lower Brominated Diphenyl Ethers – Dermal .......................... 241 3-5. Genotoxicity of Polybrominated Diphenyl Ethers (PBDEs) In Vivo................................................ 245 3-6. Genotoxicity of Polybrominated Diphenyl Ethers (PBDEs) In Vitro ............................................... 247 3-7. Ongoing Research for PBDEs in Humans ........................................................................................ 337 4-1. Chemical Identity of Polybrominated Diphenyl Ether (PBDE) Congeners ..................................... 345 4-2. Chemical Identity of Technical Polybrominated Diphenyl Ethers (PBDEs) .................................... 351 4-3. Physical and Chemical Properties of Technical Polybrominated Diphenyl Ether (PBDE) Mixtures ........................................................................................................................................... 353 4-4. Physical and Chemical Properties of Some Polybrominated Diphenyl Ether (PBDE) Congeners ......................................................................................................................................... 355 4-5. Typical Congener Composition of Penta-, Octa-, and Deca- Brominated Diphenyl Ether (BDE) ...................................................................................................................................... 357 5-1. Facilities that Produce, Process, or Use Decabromodiphenyl Ether ................................................. 360 6-1. Releases to the Environment from Facilities that Produce, Process, or Use Decabromodiphenyl Ether ................................................................................................................................................. 368 6-2. Concentrations (pg/m3) of Several Polybrominated Diphenyl Ethers (PBDEs) in Air Samples ...... 383 6-3. Concentrations (ng/g) of Several Polybrominated Diphenyl Ethers (PBDEs) in Sediment and Suspended Particulate Samples ........................................................................................................ 388 6-4. PBDE Concentrations (pg/g Wet Weight) in 18 U.S. Meat and Fish Samples ................................ 392 6-5. Concentrations (ng/g Dry Weight) of Several Polybrominated Diphenyl Ethers (PBDEs) in Biosolids (Sewage Sludge) and Effluents ........................................................................................ 396 6-6. Concentrations (ng/g Lipid Weight, Except as Noted) of Several Polybrominated Diphenyl Ethers (PBDEs) in Freshwater Fish Samples from the United States .............................................. 398 6-7. Concentrations (ng/g Lipid Weight Except as Noted) of Several Polybrominated Diphenyl Ethers (PBDEs) in Freshwater Fish Samples from Europe .............................................................. 404 PBDEs xxi 6-8. C oncentrations (ng/g Lipid Weight) of Several Polybrominated Diphenyl Ethers (PBDEs) in Marine Aquatic Species.................................................................................................................... 408 6-9. Concentrations (ng/g Lipid Weight) of Several Polybrominated Diphenyl Ethers (PBDEs) in Marine Animals ................................................................................................................................ 413 6-10. Concentrations (ng/g Lipid Weight) of Several Polybrominated Diphenyl Ethers (PBDEs) in Human Blood Samples ..................................................................................................................... 422 6-11. Concentrations (ng/g Lipid Weight) of Several Polybrominated Diphenyl Ethers (PBDEs) in Human Adipose Tissue Samples ...................................................................................................... 424 6-12. Concentrations (ng/g Lipid Weight) of Several Polybrominated Diphenyl Ethers (PBDEs) in Human Breast Milk Samples ............................................................................................................ 425 6-13. Concentrations of PBDE Congeners in Human Milk from Nursing Mothers in the United States (2001–2004) (ng/g Lipid) ...................................................................................................... 430 7-1. A nalytical Methods for Determining Polybrominated Diphenyl Ethers (PBDEs) in Biological Materials ........................................................................................................................................... 448 7-2. A nalytical Methods for Determining Polybrominated Diphenyl Ethers (PBDEs) in Environmental Samples .................................................................................................................... 451 8-1. R egulations, Advisories, and Guidelines Applicable to PBDEs ...................................................... 460 PBDEs xxii 1PBDEs1. PUBLIC HEALTH STATEMENT FOR POLYBROMINATED DIPHENYL ETHERS (PBDEs)This Public Health Statement summarizes the Division of Toxicology and Human Health Science’sfindings on PBDEs, tells you about them, the effects of exposure, and describes what you can do to limitthat exposure.The U.S. Environmental Protection Agency (EPA) identifies the most serious hazardous waste sites in thenation. These sites make up the National Priorities List (NPL) and are sites targeted for long-term federalclean-up activities. U.S. EPA has not found PBDEs in any of the 1,832 current or former NPL sites. The total number of NPL sites evaluated for PBDEs is not known. But the possibility remains that as moresites are evaluated, sites with PBDEs may be identified. This information is important because these future sites may be sources of exposure, and exposure to PBDEs may be harmful.If you are exposed to PBDEs, many factors determine whether you’ll be harmed. These include howmuch you are exposed to (dose), how long you are exposed (duration), and how you are exposed (route ofexposure). You must also consider the other chemicals you are exposed to and your age, sex, diet, familytraits, lifestyle, and state of health.WHAT ARE PBDEs?PBDEs are flame-retardant chemicals that were added to a variety of consumer products to make themdifficult to burn. These substances are not single chemical compounds, but rather mixtures of severalbrominated substances. The entire family of PBDEs consists of 209 possible substances that are referredto as congeners. There were three important commercial PBDE mixtures (i.e., penta-, octa-, and deca- bromodiphenylethers [BDEs]). DecaBDE’s main use was for electronic enclosures, such as television cabinets.OctaBDE was largely used in plastics for business equipment. PentaBDE was principally used in foamfor cushioning in upholstery. PentaBDE and octaBDE mixtures were voluntarily withdrawn from theU.S. marketplace by their manufacturers at the end of 2004 and decaBDE was not to be manufactured orimported into the United States after December 31, 2013. In 2003, the European Union (EU) passed aDirective to ban the marketing and use of penta- and octaBDE that took effect in 2004. In 2008, the useof decaBDE was restricted by a EU’s Restriction of Hazardous Substances (RoHS) Directive. 2PBDEs1. PUBLIC HEALTH STATEMENTWHAT HAPPENS TO PBDEs WHEN THEY ENTER THE ENVIRONMENT?PBDEs can be released into the air, water, and soil at places where they are produced or used. Despite thephase out of penta-, octa-, and decaBDE, vast amounts of consumer products still contain PBDEs, and these products are intended to be used for several more years. Some of these products include oldertelevisions, computers, and furniture containing polyurethane foam. PBDEs have very low water solubility, and when these substances are released to water, they typicallybind to sediment. PBDEs in consumer items put in landfills may leach through the soil into groundwater.This is not likely to be a problem, however, because these substances generally bind strongly to soilparticles, and therefore, do not move easily through soil layers.Soils and sediments are major sinks for PBDEs. Various food items, including fish, meat, and dairyproducts, have been shown to contain low concentrations of PBDEs.HOW MIGHT I BE EXPOSED TO PBDEs?Humans can be exposed to PBDEs in a wide variety of ways, including eating contaminated foods orcontaminated dusts/soils, breathing in contaminated air, or having skin contact with contaminated soil/ dust/commercial products.The primary route of exposure to PBDEs for the general population of the United States is from ingestion of contaminated dust in indoor environments, including both personal residences and work-place environments. PBDEs have been detected in residential house dust, which you can breathe in or swallowin low concentrations. This can occur because PBDEs are physically mixed into consumer products from which they have the potential to escape into the environment when conditions are ideal. Ingestion ofhouse dust (and to a lesser degree skin exposure to house dust) accounts for between 80 and 90% of totalPBDE exposures of the general population. The remaining exposure to PBDEs in the United States isfrom food ingestion. You may be exposed to PBDEs through ingestion of contaminated foods, particularly those with high fat content, such as fatty fish. In breastfeeding infants, breast milk may be a major source of PBDE exposure because PBDEs can accumulate in breast milk. Due to the chemicalnature of PBDEs, they have not been detected in water to any significant extent; therefore, drinking wateris not expected to be a major route of exposure to PBDEs. While exposure to dust appears to be the major 3PBDEs1. PUBLIC HEALTH STATEMENTexposure pathway for the general population of North American residents, PBDE exposure through dietary routes appears to be more important for European communities.PBDEs have been detected in air samples, indicating that people can also be exposed by inhalation.Consumer products such as computer and electronic equipment (e.g., televisions) treated with PBDEs can continue to release these substances to air over time.PBDEs can enter soil from discarded products (e.g., in landfills). Biosolids may also contain PBDEs;therefore, they may be inadvertently released to soils from the use of biosolids that are applied to addnutrients to farmlands. If you touch soil containing PBDEs, a small amount of PBDEs may pass throughyour skin into the bloodstream; ingestion of soil can lead to higher PBDE exposure. This route may beespecially important for children who display a lot of hand to mouth activity.HOW CAN PBDEs ENTER AND LEAVE MY BODY?PBDEs can enter your body from food, air, water, or soil. The ways that PBDEs might enter and leave your body depend on the chemical structures of the congener components. The higher-brominated PBDEs, particularly decaBDE, act somewhat differently in the body than do lower-brominated PBDEs. If you breathe air that contains PBDEs, or swallow food, water, soil, or dust contaminated with PBDEs, the lower-brominated congeners are more likely than decaBDE to enter your body through your lungs and stomach and pass into the bloodstream.Once PBDEs are in your body, the congeners might partially change into breakdown products called metabolites.PBDEs and their metabolites can leave your body, mainly in the feces and a very small amount in urine. DecaBDE, with an apparent half-time of 15 days, tends to be eliminated from your body faster than lower-brominated PBDEs, with apparent half-times as high as 94 days. Lower brominated PBDEs can stay in your body for many years, stored mainly in body fat. DecaBDE also accumulates in body fat, butto a lesser degree. Both lower-brominated PBDEs and decaBDE can concentrate in breast milk fat andcan enter the bodies of children through breastfeeding. Lower-brominated PBDEs and decaBDE also canenter the bodies of unborn babies through the placenta. 4PBDEs1. PUBLIC HEALTH STATEMENTHOW CAN PBDEs AFFECT MY HEALTH? Nothing definite is known about the health effects of PBDEs in people. The majority of information regarding toxicity of PBDEs and their breakdown products (metabolites) is from animal studies; however,several recent studies have evaluated associations between PBDE concentrations in human tissues (e.g., blood, breast milk) and various health effects. Due to differences in how decaBDE is absorbed and stored in your body, decaBDE is expected to be less toxic than lower-brominated PBDEs. Rats and mice that ingested small amounts of lower-brominated PBDEs during early development hadneurobehavioral changes and damage to their reproductive systems as adults. Altered neurobehavior was also observed in rats and mice that ingested decaBDE during early development, but at doses higher than observed for lower-brominated PBDEs. Adult rats and mice that ingested moderate amounts of lowerbrominated PBDEs for short periods of time had mainly thyroid and liver effects. Additional findingsfrom short-term animal studies suggest that some PBDEs might impair the immune system. Animalsexposed to PBDEs by skin contact showed signs of skin irritation only if they had been scratched.As with short-term exposure, rats and mice that ingested PBDEs for longer periods during earlydevelopment also showed neurobehavioral changes; again, effects occurred at higher doses withdecaBDE. Evidence from human studies is also suggestive of an association between PBDE exposure and altered neurodevelopment. Adult rats and mice that ingested small amounts of lower-brominated PBDEs over several weeks or months developed effects in the male reproductive system, thyroid, and liver. Adult animals that ingested small amounts of decaBDE over several weeks or months developedeffects in the pancreas (diabetes), nervous system, immune system, and reproductive system. Evidencefor PBDE-mediated effects from human studies in systems other than the developing nervous system is inconclusive or non-existent.We don’t know if PBDEs can cause cancer in people, although liver tumors developed in rats and micethat ate extremely large amounts of decaBDE throughout their lifetime. Lower-brominated PBDEs havenot yet been tested for cancer in animals. The International Agency for Research on Cancer (IARC) has classified PBDE as a Group 3 carcinogen (not classifiable as to its carcinogenicity to humans) based on inadequate evidence of carcinogenicity inhumans and inadequate or limited evidence in experimental animals. The EPA assigns the cancercategory Group D (not classifiable as to human carcinogenicity) to mono-, di-, tri-, tetra-, penta-, hexa-, 5PBDEs1. PUBLIC HEALTH STATEMENTocta-, and nonaBDEs and reports “inadequate information” to classify the specific congeners2,2’,4,4’-tetraBDE, 2,2’,4,4’,5-pentaBDE, and 2,2’,4,4’,5,5’-hexaBDE. However, EPA assigns a classification of “suggestive evidence of carcinogenic potential” for decaBDE. The Department of Health and Human Services has not classified PBDEs as carcinogens. The American Conference ofGovernmental Industrial Hygienists (ACGIH) has no data regarding cancer classifications for PBDEs.See Chapters 2 and 3 for more information on health effects of PBDEs.HOW CAN PBDEs AFFECT CHILDREN?This section discusses potential health effects of PBDE exposure in humans from when they’re firstconceived to 18 years of age.Studies indicate that infants and toddlers have higher exposures to PBDEs compared to older children oradults. The most likely way that infants might be exposed to PBDEs is from breast milk containingPBDEs, although fetuses in the womb could also be exposed. Toddlers and older children are exposed to PBDEs in generally the same way as are adults, mainly by ingesting contaminated household dust and food. However, soil/dust ingestion in small children (age 1−5 years) is much higher than in older children and adults. Because of their smaller weight, children’s intake of PBDEs per kilogram (or pound) of bodyweight may be greater than that of adults. Children who live near hazardous waste sites mightaccidentally eat some PBDEs by putting dirty hands or other soil/dirt covered objects in their mouths, orthrough eating without washing their hands. Some children also eat dirt on purpose. It is also possiblethat children could be exposed to PBDEs following transport of the chemical on clothing from theparent’s workplace to the home. As indicated above, young children can be exposed to PBDEs both before birth and from breast milk. Both lower-brominated PBDEs and decaBDE have been found in breast milk, and they can be transferredto babies and young children. In general, however, any risks from exposures in mother’s milk areoutweighed by the benefits of breastfeeding. You should consult your health care provider if you have any concerns about PBDEs and breastfeeding. Since the fetus and child are still developing, effects ofPBDEs might be more significant if exposure occurs during the periods before and soon after birth.Evidence suggests that fetuses and young children are more susceptible to PBDEs than adults. Subtle behavioral changes have been observed in animals exposed to PBDEs within the first 2 weeks of life, and results from human studies are suggestive of an effect of PBDEs on neurodevelopment in children, 6PBDEs1. PUBLIC HEALTH STATEMENTincluding impaired cognitive development (comprehension, memory), impaired motor skills, increased impulsivity, and decreased attention. One study reported that early PBDE exposure was a risk factor forthe development of Attention Deficit Hyperactivity Disorder (ADHD); however, another study did notfind a link between PBDE exposure and ADHD. One study investigating potential associations betweenearly PBDE exposure and autism did not find a link between maternal PBDE serum levels and autistic behaviors in 4–5-year old children. One possible explanation for the observed behavioral effects might berelated to changes in the thyroid, because development of the nervous system is dependent on thyroid hormones. Damage to developing reproductive organs and immune suppression have also been observed in animals exposed to PBDEs during development. It is unknown if these effects occur in human children.PBDEs have not caused birth defects in animals or impaired the ability for rats or mice to become pregnant or stay pregnant.HOW CAN FAMILIES REDUCE THE RISK OF EXPOSURE TO PBDEs?If your doctor finds that you have been exposed to significant amounts of PBDEs, ask whether yourchildren might also be exposed. Your doctor might need to ask your state health department to investigate.Ingestion and dermal contact with indoor dust containing PBDEs is the major exposure pathway to residents of the United States. Dust containing PBDEs can collect on your hands and be ingested throughhand-to-mouth activities; regular hand washing may decrease PBDE exposure from this route.Additionally, PBDE exposure may be decreased by regular vacuuming and cleaning of air ducts and filters to reduce indoor dust levels.Since many older consumer products such as televisions, computers, and furniture containingpolyurethane foam contain PBDEs, replacing older products with newer ones that do not contain thesesubstances may decrease residential PBDE exposure.ARE THERE MEDICAL TESTS TO DETERMINE WHETHER I HAVE BEEN EXPOSED TOPBDEs? PBDEs and their breakdown products (metabolites) can be measured in human blood, hair, and breastmilk. However, the detection of PBDEs or their metabolites cannot predict the kind of health effects that 7PBDEs1. PUBLIC HEALTH STATEMENTmight develop from that exposure. Because PBDEs and their metabolites either leave the body or aredistributed to body fat fairly rapidly, the tests need to be conducted within days if an acute, high-levelexposure is suspected. For more information on the different substances formed by PBDE breakdown and on tests to detect these substances in the body, see Chapters 3 and 7.WHAT RECOMMENDATIONS HAS THE FEDERAL GOVERNMENT MADE TO PROTECTHUMAN HEALTH?The federal government develops regulations and recommendations to protect public health. Regulationscan be enforced by law. Federal agencies that develop regulations for toxic substances include theEnvironmental Protection Agency (EPA), the Occupational Safety and Health Administration (OSHA),and the Food and Drug Administration (FDA). Recommendations provide valuable guidelines to protectpublic health but cannot be enforced by law. Federal organizations that develop recommendations fortoxic substances include the Agency for Toxic Substances and Disease Registry (ATSDR) and the National Institute for Occupational Safety and Health (NIOSH).Regulations and recommendations can be expressed as “not-to-exceed” levels; that is, levels of a toxic substance in air, water, soil, or food that do not exceed a critical value usually based on levels that affectanimals; levels are then adjusted to help protect humans. Sometimes these not-to-exceed levels differamong federal organizations. Different organizations use different exposure times (an 8-hour workday ora 24-hour day), different animal studies, or emphasize some factors over others, depending on theirmission.Recommendations and regulations are also updated periodically as more information becomes available.For the most current information, check with the federal agency or organization that issued the regulationor recommendation.EPA requires that companies that transport, store, or dispose of monobrominated diphenyl ether(monoBDE) (or diphenyl ether with one bromine attached to the structure, represented by ChemicalAbstracts Service [CAS] Registry Number 101-55-3; Resource Conservation and Recovery Act [RCRA]waste number U030) follow the rules and regulations of the federal hazardous waste managementprogram because it has been listed (U-list) as a hazardous waste due to toxicity concerns. EPA also limitsthe amount of monoBDE put into publicly owned waste water treatment plants. To minimize exposure of 8PBDEs1. PUBLIC HEALTH STATEMENTpeople to monoBDE, EPA requires that industry tell the National Response Center each time 100 poundsor more of monoBDE have been released to the environment.OSHA has not set permissible exposure limits (PELs) to protect workers against adverse health effects resulting from exposure to PBDEs. NIOSH has not recommended guidelines for worker exposure limits.WHERE CAN I GET MORE INFORMATION?If you have any questions or concerns, please contact your community or state health or environmentalquality department, or contact ATSDR at the address and phone number below. ATSDR can also providepublically available information regarding medical specialists with expertise and experience recognizing,evaluating, treating, and managing patients exposed to hazardous substances.• Call the toll-free information and technical assistance number at1-800-CDCINFO (1-800-232-4636) or• Write to:Agency for Toxic Substances and Disease RegistryDivision of Toxicology and Human Health Sciences1600 Clifton Road NEMailstop F-57Atlanta, GA 30329-4027Toxicological profiles and other information are available on ATSDR’s web site:http://www.atsdr.cdc.gov. 9PBDEs2. RELEVANCE TO PUBLIC HEALTH2.1 BACKGROUND AND ENVIRONMENTAL EXPOSURES TO PBDEs IN THE UNITED STATES PBDEs are classes of structurally similar brominated hydrocarbons in which 2–10 bromine atoms are attached to the molecular structure (i.e., diphenyl ether). Monobrominated structures (i.e., one bromineatom attached to the molecule) are often included when describing PBDEs. There are 209 differentmolecular combinations, or congeners, that are possible for PBDEs, although only a limited number exist in commercial mixtures. Based on the number of bromine substituents, there are 10 homologous groupsof PBDE congeners (monobrominated through decabrominated), with each homologous group containingone or more isomers. The mono-, di-, tri-, tetra-, penta-, hexa-, hepta-, octa-, nona-, and decabromocongeners can exist in 3, 12, 24, 42, 46, 42, 24, 12, 3, and 1 isomers, respectively. The general chemicalstructure of PBDEs is shown below:2 2'O BrnBrm 1' 3' 4' 5' 6' 13 4 5 6 where m + n = 1 to 10Due to the ether linkage and the position and number of bromine atoms, there are important three-dimensional differences in the structures of PBDEs that can influence the molecules’ receptor interactions and toxicological properties as discussed in Section 3.5, Mechanisms of Action. In general, PBDEs arenot expected to have the same array of three-dimensional conformations as either polybrominated biphenyls (PBBs) or polychlorinated biphenyls (PCBs).PBDEs are brominated organic compounds that were used as flame retardant additives in plastics, textiles,and other materials. As additives, they are physically mixed into product applications, rather than chemically bound. Therefore, they have the potential to migrate from the plastic matrix into theenvironment when conditions are ideal. Production of PBDEs began in the 1970s and has continued untilrecently. PentaBDE and octaBDE mixtures were voluntarily withdrawn from the U.S. marketplace bytheir manufacturers at the end of 2004; however, the manufacture and use of decaBDE continued past thatdate. In December of 2009, the two remaining U.S. producers of decaBDE and the largest U.S. importer PBDEs 102. RELEVANCE TO PUBLIC HEALTHof this product announced commitments to phase out manufacture and importation of decaBDE for mostuses in the United States by December 31, 2012, and to end manufacture and import for all uses by theend of 2013.PBDEs are persistent in the environment and most congeners can be considered bioaccumulative. Thefully brominated congener, BDE 209, has a lower tendency to bioconcentrate due to its moleculardimensions; however, numerous studies have detected this substance in tissues of birds, mammals, andfish (see Section 6.4.4). Monitoring and body-burden data indicate that PBDEs are ubiquitous in the environment and that thegeneral population is exposed to these substances through their past use as flame retardants. A study thatexamined stored blood samples from 1973 (prior to the use of PBDEs as flame retardants) showed virtually no detections of these substances in human blood; however, varying concentrations of manyPBDE congeners were detected in all blood samples collected in 2003 from 39 residents in Mississippiand 10 residents in New York City, illustrating the widespread exposure to PBDEs since their inception. Body burden data have consistently shown the residents of North America have higher concentrations ofPBDEs in blood than people residing in Europe, likely due to differences in past production and use ofcommercial formulas.House dust was identified as a major source of exposure to PBDEs by a systematic study of Americanexposure routes. In the United States, concentration levels found in soil, house dust, and air tended to begreater in indoor samples compared to outdoor samples. The EPA calculated the adult intake dose of totalPBDEs to be 7.1 ng/kg body weight/day. It estimated children’s intakes as 47.2 ng/kg body weight/dayfor 1–5 year olds, 13.0 ng/kg body weight/day for 6–11 year olds, and 8.3 ng/kg body weight/day for 12– 19 year olds. The much higher dose for children aged 1−5 years was largely due to higher soil/dustingestion in this age group. Exposure to indoor dust was the predominant exposure pathway for PBDEsin these calculations. It was estimated that 90% of the intake resulted from house dust inhalation ordermal exposure. This does not include special populations such as infants that are primarily exposed through breastfeeding. PBDE concentrations were generally lower in house dust samples collectedoutside of the United States compared with dust samples collected within the United States. The specific PBDE congeners detected in house dust and food vary; BDE 209 is more commonly detected within indoor environments where exposure is more likely to occur through intake of contaminated dust and air. PBDE contamination of food is more likely a result of past emissions or ongoing emissions fromdumpsites and older products that still contain pentaBDE, which is mainly composed of the congeners PBDEs 112. RELEVANCE TO PUBLIC HEALTHBDE 47 and BDE 99. Ingestion of PBDEs through the diet appears to be the predominant pathway forEuropean communities. In China, the decaBDE congener, BDE 209, was the most abundant congenerdetected in both maternal and cord blood samples where industrial production of BDE 209 may result in exposure.2.2 SUMMARY OF HEALTH EFFECTS Information is available on the potential health effects of formerly used commercial PBDE mixtures (pentaBDE, octaBDE, and decaBDE) as well as several individual PBDE congeners. As subsequentlydiscussed, the toxicity of decaBDE is generally less pronounced than for lower-brominated PBDEsfollowing acute and repeated-dose exposures. This difference in toxicity may be related to differences inpharmacokinetics, resulting in lower bioavailability of decaBDE (see Sections 2.3 and 3.4 for moredetails). The preponderance of health effects information on PBDEs is from studies of orally exposed laboratoryanimals and human studies in which the main exposure route is unknown, but expected to be oral. As summarized below and detailed in Chapter 3 (Health Effects), the main targets of concern followingPBDE exposure in humans are the developing nervous and reproductive systems, the developing and mature endocrine system, the liver, and the male reproductive system. Other potential targets are the female reproductive system, the adult nervous system, and the developing and adult immune system;however, evidence for these end points is limited. In other systems/organs, available data provide noconsistent evidence for exposure-related effects (respiratory, cardiovascular, gastrointestinal,hematological, renal, dermal, or ocular effects); therefore, effects in these systems are considered unlikely to occur following PBDE exposure. Developmental Effects.Neurodevelopment. Numerous epidemiological studies have reported results suggestive of an effect ofPBDE on neurodevelopment in children. PBDE concentrations in cord blood, maternal or infant serum, and/or breast milk have been correlated with cognitive score and adaptive behavior deficits in infants; mental and physical development deficits in infants/toddlers at ages 12, 24, and 36 months; language and social developmental score deficits in toddlers at 24 months; increased impulsivity in toddlers at 24– 36 months; poor social competence and attention deficit hyperactivity disorder (ADHD) or increasedattention problems in 4-year-old children; impaired fine motor coordination, verbal memory and PBDEs 122. RELEVANCE TO PUBLIC HEALTHcomprehension, and sustained attention in 5–7-year-old children; and poor attention and executivefunction deficits in 9–12-year-old children. In one birth cohort, no associations were observed between maternal serum PBDEs and neonatal behavior in 5-week-old infants or autistic behaviors at 4–5-year-old children; however, children from the same cohort showed associations between maternal serum PBDEs and decreased IQ and increased hyperactivity at 5 years of age and executive function deficits at 5– 8 years of age. Pre- and perinatal studies in animals also consistently reported neurodevelopmentaleffects following exposure to lower-brominated PBDEs and decaBDE at doses ≥0.06 and ≥2.22 mg/kg/day, respectively, including neurobehavioral alterations, delayed ontogeny of reflexes, ultrastructural changes in the hippocampus, altered nicotinic receptor density, altered electrophysiology, and altered gene and protein expression levels. Based on human and animal data, the developing nervous system is a target of concern for both lowerbrominated PBDEs and decaBDE.Endocrine System Development. In infants, developmental exposure to PBDEs and infant serum or cord blood thyroxine (T4) levels were either negatively associated or not associated. Associations between developmental PBDE exposure and infant serum or cord blood triiodothyronine (T3) and thyroid stimulating hormone (TSH) were similarly inconsistent. In animals, numerous studies have reported decreased serum T4 and/or T3 levels in pups after gestational and lactational exposure to penta- ortetraBDE at doses as low as 0.3 mg/kg/day in rats and 452 mg/kg/day in mice. Significant reductions in serum T3 levels were also observed in offspring following gestational and lactational exposure to decaBDE at 146 mg/kg/day in rats and 1,500 mg/kg/day in mice, although no changes were observed in serum T4 or TSH. A dose-related decrease in serum T4 was observed in neonatal male mice exposed todecaBDE doses of 6–20 mg/kg/day from postnatal day (PND) 2 to 15, but no change was observed in neonatal females. While human data are inconsistent, they suggest that PBDEs can interact with thyroid hormonehomeostasis in infants and children. These data, along with available animal studies, indicate that the developing thyroid is a target of concern for PBDE exposure, especially lower-brominated PBDEs. Reproductive System Development. Male reproductive effects significantly associated with PBDEexposure in infants included congenital cryptorchidism (undescended testes), decreased cord serum total testosterone (but not free testosterone, estradiol [E2], aromatase index, sex hormone binding globulin, orAnti-Müllerian hormone), increased serum levels of the sex hormones, E2, free E2, and inhibin B (but not PBDEs 132. RELEVANCE TO PUBLIC HEALTHtestosterone, luteinizing hormone [LH], follicle-stimulating hormone [FSH], or sex hormone bindingglobulin) at 3 months of age, and increased testes volume in boys at 18 months of age. In contrast, no relationships were observed between maternal PBDE exposure and hypospadias (abnormal location of theurinary tract opening) in male offspring, PBDE concentrations in children’s adipose tissue and cryptorchidism, or various measures of sexual maturation in female offspring. However, serum PBDElevels in 6–8-year-old females were significantly associated with delayed onset of puberty in alongitudinal cohort of U.S. girls.In animal studies, reproductive effects were observed in adult F1 offspring of dams exposed to a single dose of pentaBDE at 0.06 mg/kg on gestation day (GD) 6, including reductions in testicular weight, sperm/spermatid number, and daily sperm production in males and a decreased number of secondaryfollicles and ultrastructural changes in the ovaries in females (although F1 fertility when mated to anunexposed animal was not impaired). In animals exposed pre- or perinatally to decaBDE, reproductiveeffects were observed in adult male offspring of dams exposed to decaBDE doses of 10–1,500 mg/kg/day from GD 0 to 17, including testicular lesions, decreased anogenital distance (AGD), and altered spermparameters. In contrast, no exposure-related changes in AGD, onset of puberty, or reproductive organ weight and histology were reported in offspring of dams exposed to decaBDE at doses up to 1,000 mg/kg/day during gestation and lactation or doses up to 20 mg/kg/day from PND 2 to 15. No exposure-related changes in reproductive development were reported in female offspring of dams exposedto decaBDE at doses up to 1,000 mg/kg/day during gestation and lactation or doses up to 20 mg/kg/dayfrom PND 2 to 15.Based on limited human and adequate animal data, it is possible that oral PBDE exposure duringdevelopment may adversely affect the developing reproductive system, particularly the male reproductive system. However, data are too limited to adequately determine whether or not PBDE exposure in infantsand children will lead to altered reproductive performance as adults.Immune System Development. In offspring of rat dams exposed to pentaBDE at doses up to 25 mg/kg/day via gavage for 70 days prior to mating through PND 21, a significant dose-related trendwas observed in the incidence of apoptotic lymphocytes and tingible macrophages in the thymus ofPND 43 males, but not females. In PND 28 offspring of mouse dams exposed to decaBDE at doses≥260 mg/kg/day from GD 10 to PND 21, pulmonary viral titers of respiratory syncytial virus (RSV)(measured 5 days post-infection) were significantly increased. These animal data suggest that oral PBDE PBDEs 142. RELEVANCE TO PUBLIC HEALTHexposure during development may lead to immunosuppression; however, data are too limited to adequately assess the immunotoxic potential of PBDE exposure in infants and children.Embryotoxicity, Fetotoxicity, and Physical Growth and Development. No human studies have evaluatedassociations between embryotoxicity or fetotoxicity and PBDE exposure. One study reported asignificant increased risk for preterm birth in women with high serum PBDE concentrations, compared with low PBDE concentrations; however, other studies did not observe a significant association betweengestational length and PBDE concentrations. Evidence for altered physical growth and development fromhuman studies is inconsistent, with some studies reporting associations between PBDE concentrations in maternal/cord serum, breast milk, or placental tissue and decreased birth weight, length, chestcircumference, head circumference, and/or body mass index (BMI), some reporting no associations, and alimited number reporting increased birth weight, length, and/or head circumference. Available data fromanimal studies indicate that PBDEs are not embryotoxic or fetotoxic at PBDE doses below doses thatelicited maternal toxicity, although occasional observations of reduced pup weight were reported.Based on animal studies, it is unlikely that oral PBDE exposure will cause embryotoxicity or fetotoxicityin humans; however, human and animal data indicate that PBDE exposure could potentially lead to lowbirth weight or other impacts on physical development.Endocrine Effects.Thyroid. Numerous studies have been performed to evaluate the relationship between concentrations ofPBDE in body tissues and circulating thyroid hormone levels in human populations. While these studieshave demonstrated that PBDE can perturb the human endocrine system and affect hormone levels, thespecific findings are not consistent across studies. For example, even limiting the discussion to studiesthat evaluated both PBDE concentrations and thyroid hormone levels in serum samples collected onlyfrom adult men, studies have reported positive associations with T4, negative associations with T4, and no association with T4. Similar inconsistencies were found in studies in pregnant women, with studiesreporting a positive associations with T4, no association with T4, or negative associations with T4. Resultswere similarly inconsistent regarding the association between PBDE concentrations and serum T3 and TSH.In contrast to inconsistencies observed in human studies, altered serum thyroid hormone levels have been consistently reported in laboratory animals exposed to lower-brominated PBDEs. Reduced serum T4 has PBDEs 152. RELEVANCE TO PUBLIC HEALTHbeen reported in animals following acute or intermediate exposure to lower-brominated PBDEs at doses as low as 0.8 mg/kg/day. At higher doses (≥30 mg/kg/day), some studies also report reduced serum T3 and/or increased serum TSH; however, other studies reported no significant changes in serum T3 and/orTSH levels in rats exposed to doses up to 300 mg/kg/day. In rat dams, reduced serum T4 has beenobserved following exposure to lower-brominated PBDEs at doses as low as 0.06 mg/kg/day duringgestation or gestation plus lactation. In mouse dams, no exposure-related changes were observedfollowing exposure to pentaBDE at 452 mg/kg/day from GD 4 to PND 17. Exposure to pentaBDE atdoses up to 120 mg/kg/day during gestation or gestation plus lactation did not significantly alter maternal serum T3 and/or TSH in most studies; however, a study reported reduced maternal serum T3 afterexposure to pentaBDE at 30 mg/kg/day via dosed cookies from GD 1 to PND 21.Histopathological changes in the thyroid (e.g., follicular cell hyperplasia, increased epithelialthickness/height, altered morphology of epithelium, cellular debris, degeneration) have been observed in intermediate-duration studies of lower-brominated BDEs at doses as low as 20 mg/kg/day in rats,0.45 mg/kg/day in mice, and 0.06 mg/kg/day in mink. Increased thyroid weights were reported in ratsexposed to doses of penta- or octaBDE as low as 50 mg/kg/day for 90 days; however, no exposure-relatedchanges in thyroid weights were observed in rats exposed to doses of penta- or octaBDE up to 200 mg/kg/day for 15–28 days or in F0 or F1 mink exposed to dietary pentaBDE at doses up to 0.31 mg/kg/day in one-generation studies (4 weeks pre-mating through postnatal week [PNW] 6 or 33). In acute studies, no exposure-related changes in thyroid weight or histology were observed in rats exposed to penta- or tetraBDE doses up to 36 mg/kg/day for 14 daysUnlike the lower-brominated PBDEs, serum T4 levels were not altered in rats exposed to decaBDE atdoses up to 600 mg/kg/day for 4–90 days. Some studies reported reduced serum T3 levels in rats exposedto decaBDE at doses as low as 50 mg/kg/day for 28–90 days, but another 90-day study reported no change in serum T3 levels in rats exposed to decaBDE at doses up to 100 mg/kg/day. Serum TSH was reduced in male rats exposed to decaBDE at doses ≥300 mg/kg/day for 33 days. At higher doses(≥950 mg/kg/day), significant reductions in serum T4 and T3 were observed in male mice exposed for35 days and pregnant mice exposed from GD 7 to 9. In chronic studies of decaBDE, thyroid follicular cell hyperplasia was observed in male mice exposed to≥3,200 mg/kg/day for 103 weeks; no histopathological changes in the thyroid were observed at doses up to 7,780 mg/kg/day in female mice, 2,240 mg/kg/day in male rats, or 2,550 mg/kg/day in female rats. Inintermediate-duration studies, dose-related increases in thyroid hyperplasia were reported for male rats PBDEs 162. RELEVANCE TO PUBLIC HEALTHexposed to a low-purity decaBDE compound at ≥80 mg/kg/day for 30 days, but hyperplasia was notobserved in rats or mice exposed to high-purity decaBDE at doses up 8,000 or 9,500 mg/kg/day,respectively, for 13 weeks or in rat dams exposed to doses up to 146 mg/kg/day from GD 10 to PND 21. However, multiple areas of degenerated follicular epithelium and slight attenuation of the follicularepithelium were observed in the thyroid glands of young male rats exposed to decaBDE at doses≥300 mg/kg/day for 33 days. No changes in thyroid weight were observed in rats exposed to decaBDE atdoses up to 90 mg/kg/day for 28 days, but increased thyroid weights were reported in rat dams exposed to ≥2 mg/kg/day from GD 10 to PND 21 and young male rats exposed to 600 mg/kg/day for 33 days.While human data are inconsistent, they suggest that PBDEs can interact with thyroid hormonehomeostasis. These data, along with available animal studies, indicate that the thyroid is a target ofconcern for PBDE exposure, especially lower-brominated PBDEs.Pancreas. An analysis of cross-sectional National Health and Nutrition Examination Survey (NHANES)data showed a significant increase in the risk of diabetes associated with serum concentrations of BDE153 (but not BDE 28, BDE 47, BDE 99, or BDE 100), although the risk was higher with exposure to 50– 75th percentile BDE 153 concentrations than >75th percentile BDE 153 concentrations. Serum BDE 153 concentrations (but not BDE 28, BDE 47, BDE 85, BDE 99, BDE 100, or BDE 154) were also shown to be significantly associated with increased odds of developing gestational diabetes in a cohort of258 pregnant women. However, other cross-sectional and prospective studies found no relationship between serum PBDE concentrations and diabetes in an adult cohort from Wisconsin, an elderly cohort in Finland, or an elderly cohort in Sweden.An intermediate-duration study evaluated insulin-regulation and pancreatic morphology in rats following exposure to decaBDE at 0, 0.05, 1, or 20 mg/kg/day daily via gavage in corn oil for 8 weeks. Seruminsulin was significantly decreased by 50–60% at 1 and 20 mg/kg/day, and glucose levels wereconcomitantly increased by 12, 18, and 21% at 0.05, 1, and 20 mg/kg/day. Consistent with the insulin findings, morphological changes were seen in the pancreas at 1 and 20 mg/kg/day (blurred boundariesamong pancreatic islet cells; quantitative data not reported). Additionally, microarray analysis indicatedthat type I diabetes mellitus (T1DM) canonical pathways were significantly enriched following decaBDEexposure. Subsequently, gene act network and gene coexpression network found that some majorhistocompatibility complex molecules and TNF-α were involved in the T1DM pathway. Only one otheranimal study evaluated the pancreas following decaBDE exposure. In rats exposed to decaBDE viagavage for 28 days at doses of 0, 0.27, 0.82, 2.47, 7.4, 22.2, 66.7, or 200 mg/kg/day, slight or moderate PBDEs 172. RELEVANCE TO PUBLIC HEALTHinsulitis was observed in the Langerhan’s islets of the “majority of samples,” but findings were notexposure-related. Similarly, no exposure-related effects were observed for serum glucose levels. Theonly other study evaluating serum glucose levels after decaBDE exposure instead reported reduced serumglucose levels in male rats exposed to 20 mg/kg/day of a dietary PBDE mixture containing 52.1%pentaBDE (DE-71), 44.2% decaBDE (BDE 209), and 0.4% octaBDE (DE-79) for 70 days. The observed decreased glucose levels could be due to the pentaBDE component, as male rats exposed to pentaBDE atdoses of 0.27–200 mg/kg/day for 28 days also showed decreased glucose levels; the study authors did notreport the lowest dose at which glucose levels were significantly lower in male rats, but they reported a BMD10RD of 179.55 mg/kg/day and a BMDL10RD of 66.7 mg/kg/day. Limited human evidence is inconclusive regarding potential associations between diabetes and PBDEexposure; however, considering the animal data, the pancreas may be a target of concern for oral PBDEexposure. Hepatic Effects. Liver effects have been reported in adult, pregnant, and developing animals exposed to lower-brominated PBDEs. Histopathological effects in the liver (hepatocellular hypertrophy, necrosis, and vacuolation) were consistently observed in animals exposed to lower-brominated PBDEs for 15– 90 days at doses ≥0.45 mg/kg/day, regardless of life-stage. In acute exposure studies, fatty degenerationof the liver was observed following a single pentaBDE dose of 2,000 mg/kg or repeated pentaBDE doses of 200 mg/kg/day for 7 or 14 days. Increased liver weight was also observed in animals exposed tolower-brominated PBDEs for 1–14 days at doses ≥8 mg/kg/day and 15–90 days at doses ≥1.2 mg/kg/day. In studies that evaluated hepatic enzyme induction (e.g., elevated ethoxyresorufin o-deethylase [EROD],methoxyresorufin o-deethylase [MROD], pentoxy-resorufin o-deethylase [PROD], and uridinediphosphoglucuronyl transferase [UDPGT] activity), significantly increased enzyme activities were observed in animals following acute or intermediate-duration exposure to doses ≥3 or ≥0.06 mg/kg/day,respectively, and were always observed at doses at or below the dose causing elevated liver weights in the same study. No studies evaluating liver effects following chronic exposure to lower-brominated PBDEswere located.Evidence for hepatic toxicity following exposure to decaBDE is less consistent than evidence for lowerbrominated PBDEs. There is no evidence of hepatic toxicity following acute exposure to decaBDE atdoses up to 1,000 mg/kg/day for 4–14 days. In intermediate-duration studies, slight to moderatehepatocellular hypertrophy was observed in rats exposed to decaBDE at 60 mg/kg/day for 28 days, in pregnant rats exposed to 300 mg/kg/day for 21 days, and in mice exposed to 9,400 mg/kg/day for PBDEs 182. RELEVANCE TO PUBLIC HEALTH28 days; however, other studies did not report exposure-related changes in liver histology followingexposure to decaBDE at doses up to 9,500 mg/kg/day for 3–13 weeks. In an older study using an impuredecaBDE compound containing lower-brominated congeners (77% decaBDE, 22% nonaBDE, 0.8%octaBDE), centrilobular cytoplasmic enlargement and vacuolation, as well as increased liver weight, wereobserved in male rats exposed to dietary doses of ≥80 and 800 mg/kg/day, respectively, for 30 days. Increased liver weights were also reported in rats exposed to doses ≥1 mg/kg/day for 8 weeks and miceexposed to 9,400 mg/kg/day for 28 days, but other intermediate-duration studies reported no exposure-related changes in liver weights at doses of 1,000 mg/kg/day for 21–90 days. In chronic studies, exposureto decaBDE for 103 weeks caused liver lesions that included neoplastic nodules in rats at≥1,120 mg/kg/day, thrombosis and degeneration in rats at 2,240 mg/kg/day, and centrilobular hypertrophyand granulomas in mice at ≥3,200 mg/kg/day.Developing animals appear to be more susceptible to liver damage following exposure to decaBDE than adult animals. Transient histopathological changes (diffuse liver cell hypertrophy with increasedcytoplasmic eosinophilia) and elevated liver weights were observed in male and female rat offspringexposed to decaBDE from GD 10 to PND 21 at ≥2 and 146 mg/kg/day, respectively. Fatty degeneration, elevated liver weights, and elevated liver enzymes were observed in young male rats exposed to decaBDEat ≥300 mg/kg/day from PND 10 to 42. In mice, decaBDE exposure from GD 0 to 17 caused acute cellswelling of hepatocytes associated with pressure occlusion of hepatic sinusoids and elevated liverenzymes in PND 71 male offspring at ≥10 mg/kg/day; however, liver weight was unaltered at doses up to600 mg/kg/day. No studies are available on hepatic effects of PBDEs in humans. Based on the evidence in animals,PBDEs are potentially hepatotoxic in humans, especially lower-brominated PBDEs.Male Reproductive Effects. Several studies have found results suggestive of reproductive effects in men associated with exposure to PBDE, including significant inverse correlations between serumconcentrations of BDE 153 (hexaBDE) and sperm concentration and testis size in young adult Japanesemales, significantly reduced sperm mobility in association with increased serum PBDE concentrations(BDE 47, BDE 100, and total) in Canadian men recruited at a fertility clinic, and altered parameters ofsemen quality associated with selected BDEs in men participating in a prospective cohort study in Michigan and Texas. Although a number of studies have evaluated the potential effects of PBDEexposure on male reproductive hormone levels, these studies collectively do not show consistent effects associated with PBDE exposure. PBDEs 192. RELEVANCE TO PUBLIC HEALTHReproductive effects have been reported in male rodents following intermediate-duration exposure to thelower-brominated PBDE congener tetraBDE, including decreased serum testosterone at≥0.001 mg/kg/day, histopathological changes in rat or mouse testes (increased epithelial thickness,multinucleated giant cells, vacuolar spaces, apoptosis, germ cell loss) at doses ≥0.03 or≥0.045 mg/kg/day, and decreased sperm production in rats at 1 mg/kg/day. No dose-related changes were observed in testicular weight or sperm morphology, motility, or capacitation at doses up to 30 mg/kg/day. For other lower-brominated PBDEs (pentaBDE, octaBDE), no exposure-related effects were observed inserum testosterone levels at doses up to 60 mg/kg/day for 15–70 days, sperm parameters at doses up to 20 mg/kg/day for 20 days, or male reproductive histology or organ weight at doses up to 750 mg/kg/day for 38–90 days. However, in acute studies, serum testosterone levels were significantly decreased by~40–45% in male rats 45 days after a single gavage exposure to pentaBDE at doses ≥0.6 mg/kg and dose-related decreases in androgen-dependent tissue weights (prostate, seminal vesicle, Cowper’s gland, glandpenis, levator ani bulbo cavernosus) were observed in castrated rats exposed to pentaBDE at doses ≥30 mg/kg/day for 9 days (Herschberger assay).Studies of decaBDE have been more limited. Decreased relative testes and epididymides weights,degenerative changes in the seminiferous tubules, reduced sperm count and viability, and reduced serumtestosterone were reported in male mice exposed to decaBDE at 950 mg/kg/day via gavage for 35 days,compared with controls; no exposure-related effects were observed at 750 mg/kg/day. In contrast, no changes in sperm count, motility, or morphology were observed in rats or mice exposed to decaBDE atdoses up to 60 or 1,500 mg/kg/day for 28–50 days; however, exposure-related decreases were observed inone sperm velocity measure (lateral head amplitude) in mice exposed to ≥500 mg/kg/day. A dose-relateddecrease in epididymis weight and a dose-related increase in seminal vesicle/coagulation gland weight were observed in rats exposed to 1.7–60 mg/kg/day for 28 days; however, the lowest doses at which the effects were observed were not reported. No testicular weight changes were observed at doses up to 60 mg/kg/day. In other intermediate-duration studies, no exposure-related changes in organ weight were reported for male reproductive organs in rats or mice exposed to decaBDE doses up to 800 or1,500 mg/kg/day. Histopathological changes in male reproductive tissues have not been reported in rats or mice exposed to decaBDE at doses up to 2,550 or 7,780 mg/kg/day, respectively, for 103 weeks. Findings were negative in a one-generation study that exposed male and female rats to an impuredecaBDE compound containing lower-brominated congeners (77% decaBDE, 22% nonaBDE, 0.8%octaBDE) for 60 days prior to mating through PND 21. PBDEs 202. RELEVANCE TO PUBLIC HEALTHBased on the evidence in humans and animals, lower-brominated PBDEs are potentially toxic to the male reproductive system in humans. Available data for decaBDE provide very limited evidence of male reproductive damage.Female Reproductive Effects. Evidence for reproductive effects in women associated withexposure to PBDEs is inconsistent. Increased length of menstrual periods (prior to pregnancy) was associated with increased breast milk concentrations of PBDEs in a study of 46 Taiwanese women, and with plasma levels of BDE 47 and BDE 153 in a study of 42 Cree women of James Bay, Canada, but notin a smaller study with 20 Taiwanese women or in 223 Californian women. Age at menarche was notassociated with PBDE concentrations in breast milk; however, an analysis of cross-sectional data from a sample of 271 adolescent girls (NHANES) found that higher serum PBDE concentrations were associatedwith younger age of menarche. Decreased fecundability (i.e., increases in time to pregnancy betweenstopping contraception and becoming pregnant) was significantly associated with increased serumconcentrations of BDE 47, BDE 99, BDE 100, and BDE 153 (and their sum) in 223 Californian women; however, fecundability was not significantly related to serum PBDE concentrations in a cohort of501 Michigan and Texas couples followed prospectively for 1 year after discontinuing contraception forthe purpose of becoming pregnant. A study of 65 women from Boston undergoing in vitro fertilizationfound no association between serum PBDE concentrations and risk of implantation failure, but did find asignificantly increased risk of failure associated with increased (i.e., above median) concentrations ofBDE 153 (but not other congeners or total PBDE) in follicular fluid. PBDEs were not associated withPolycystic Ovary Syndrome or with diagnosis of uterine fibroids. In one-generation animal studies, no exposure-related changes were observed in reproductive end points(number of pregnancies, gestation length, number, size, or sex ratio of litters) in rats or mice exposed to lower-brominated PBDEs at doses up to 25 or 1 mg/kg/day, respectively. Similarly, in gestation pluslactation studies, no exposure-related effects on litter parameters (successful delivery of litters, gestationlength, litter size, sex ratio, number of live pups) were observed in rats or mice exposed to lowerbrominated PBDEs at doses up to 32 or 10 mg/kg/day, respectively, during gestation and lactation only. The number of litters surviving until PND 8 was significantly decreased following exposure to tetraBDEat 0.1 mg/kg/day from pre-mating day 28 through PND 21 in one study; however, reduced pup survivalwas not reported in other studies. In a one-generation study in mink, females exposed to pentaBDE atdoses ≥0.25 mg/kg/day from pre-mating day 28 through PNW 6 did not whelp. It is not clear in one studywhether mink exposed to 0.25 mg/kg/day never became pregnant or had complete litter loss. However, another study reported that female mink exposed to 0.31 mg/kg/day had no exposure-related changes in PBDEs 212. RELEVANCE TO PUBLIC HEALTHmating success; rather, sows showed complete litter loss with 70% showing clear postimplantation loss.In one-generation and intermediate-duration studies, no changes in reproductive organ weight orhistology were observed in female rats, mice, or mink exposed to lower-brominated PBDEs at doses up to 750, 0.45, or 0.31 mg/kg/day, respectively. One acute study reported increased paired ovary weight afterexposure to tetraBDE at 0.14 mg/kg on GD 6; however, no changes in reproductive organ weight orhistology were observed in female rats acutely exposed to pentaBDE at doses up to 300 mg/kg/day.In female mice exposed to dietary tetraBDE for 28 days, serum testosterone and E2 were significantlyincreased at 0.45 mg/kg/day; no other study reported altered reproductive hormones in females.Information on reproductive effects of decaBDE is limited. Findings were negative in a one-generation study that exposed male and female rats to an impure decaBDE compound contaminated with lowerbrominated congeners (77% decaBDE, 22% nonaBDE, 0.8% octaBDE) for 60 days prior to mating through PND 21. In a gestational exposure study in mice (GDs 7–9), significant increases were observedin the percentage of postimplantation loss per litter and resorptions per litter in dams exposed to ≥750 and≥1,500 mg/kg/day, respectively. Additionally, the percentage of live fetuses per litter was significantly decreased by 10% in dams exposed to 2,000 mg/kg/day. Histological changes in the ovaries (atrophic changes, decreased number of follicles, and increased fibrotic tissue) were observe in female rats exposedto decaBDE at 300 mg/kg/day from 3 weeks of age, through mating to untreated males, gestation, andlactation (~11 weeks); F0 reproductive success was not reported. In other studies, no histopathologicalchanges in female reproductive organs were observed in rats or mice following intermediate- or chronic-duration exposure to decaBDE at doses up to 8,000 or 9,500 mg/kg/day, respectively.Based on inconsistent data in humans and animals, it is unclear whether PBDEs affect the female reproductive system in adults.Adult Neurological Effects. While neurobehavioral development is a potential effect of concern forPBDE exposure in humans, available human data are too limited to determine if PBDE exposure is neurotoxic in adults or adolescents. No association was found between serum PBDE concentrations and neuropsychological function assessed by 34 tests of cognitive and motor function, affective state, and olfactory function in a study population of 144 volunteers (67 males and 77 females) between the ages of55 and 74 years who lived for at least 25 years in the upper Hudson valley of New York State. In 515 secondary students from Belgium (mean age 14.9 years), serum PBDE concentrations were notassociated with most aspects of neurological performance measured in a battery of neurological tests; PBDEs 222. RELEVANCE TO PUBLIC HEALTHhowever, there was a significant deterioration in performance in the finger tapping test with increasingPBDE level, suggesting an effect of PBDE on motor activity.Evidence for neurological effects of lower-brominated PBDEs in adult animals is limited. In repeat-exposure neurobehavioral studies, impaired learning and memory were observed in male rats exposed totetraBDE at ≥0.1 mg/kg/day for 30 days and impaired attention and inhibitory control were observed in male mice exposed to pentaBDE at 26.2 mg/kg/day for 125 days. No exposure-related neurobehavioralchanges were observed in rats exposed to pentaBDE at doses ≤17.5 mg/kg/day for 90–125 days or male rats exposed once to pentaBDE doses up to 1.2 mg/kg/day. No exposure-related changes in brain weightand/or histology were observed in animals exposed to lower-brominated PBDEs at doses up to 750 mg/kg/day.Evidence for neurological effects of decaBDE in adult animals is extremely limited. Decreased anxietybehavior in the elevated-plus maze was observed in male mice exposed to decaBDE for 15 days;however, no exposure-related changes were observed in anxiety behaviors in the light/dark test, in learning or memory in the Morris water maze, or in general neurological behaviors assessed using a functional observation battery. In another study, no changes were observed in open-field behavior ofmale rats exposed to decaBDE at doses up to 50 mg/kg/day via gavage for 90 days. No changes in brain weight were observed in rats or mice exposed to decaBDE at doses up to 90 or 160 mg/kg/day, respectively, for 15–60 days. No overt signs of neurotoxicity were observed in rats and mice exposed todecaBDE in estimated dietary doses as high as 16,000–19,000 mg/kg/day for 14 days, 8,000– 9,000 mg/kg/day for 13 weeks, or 2,550–7,780 mg/kg/day for 103 weeks. Although the high doses and extended exposure durations provided opportunities for the induction and/or development of clinicalsigns, the study is limited by lack of testing for subtle behavioral changes and neurodevelopmentaleffects.Based on available data in humans and animals, it is unclear whether PBDEs affect the adult nervous system.Immunological and Lymphoreticular Effects. Limited human data regarding potentialimmunotoxic effects of PBDEs are available. A significant negative association was found betweenserum concentrations of lower-brominated PBDE and number of circulating lymphocytes in a subset of a cohort of 33 adolescent children from the Netherlands. No effects on pokeweed mitogen-stimulated DNAproliferation or IgG immunoglobulin synthesis were found in human lymphocytes exposed to lower PBDEs 232. RELEVANCE TO PUBLIC HEALTHbrominated PBDEs in vitro. Studies of Swedish subjects reported a significantly negative associationbetween serum levels of BDE 47 and levels of protein complement 3, but not with levels of multipleinflammatory markers. In a study of Chinese children, serum levels of BDE 28 and 209 were associatedwith an increased risk of asthma. There is limited evidence for impaired immune function in animals following exposure to lowerbrominated PBDEs; however, comprehensive immunological evaluations have not been performed on anycongener or previously used commercial mixture. The plaque-forming splenic cell antibody response toinjected sheep red blood cells was significantly reduced in mice exposed to 72 mg/kg/day pentaBDE for14 days; single doses as high as 500 mg/kg had no effect. In the same study, exposure to up to 72 mg/kg/day had no effect on natural killer cell (NKC) activity. In vitro production of IgGimmunoglobulin from pokeweed mitogen-stimulated splenocytes was reduced in mice exposed to36 mg/kg/day pentaBDE for 14 days. Other 14-day studies in mice found no changes in NKC activity tomurine YAC-1 target cells at pentaBDE doses up to 72 mg/kg/day or numbers of splenic and thymiclymphocyte subsets at pentaBDE doses up to 36 mg/kg/day, although 18 mg/kg/day of tetraBDE caused significantly reduced numbers of total lymphocytes and CD4+, CD8+, and CD45R+ subtypes in spleen. In the only intermediate-duration study evaluating immune function, no dose-related changes were observed in antibody-mediated immunity to keyhole limpet hemocyanin (KLH) or phytohemagglutinin (PHA) skin response in mink exposed to pentaBDE at doses up to 0.78 mg/kg/day for 9 weeks.Histopathological changes in the spleen (hyperplasia, germinal center development) were reported insome studies following intermediate-duration exposure to lower-brominated PBDEs at doses as low as 0.63 mg/kg/day in mink and at 0.45 mg/kg/day in mice. In other studies, no exposure-related changes were observed in spleen, thymus, lymph node, and/or bone marrow tissue histology in rats exposed to lower-brominated PBDEs at doses up to 750 mg/kg/day for 28–90 days. Additionally, no exposure-related changes were observed in the histology of the spleen, thymus, Peyer’s patches, or mesenteric lymph nodes in rat dams exposed to pentaBDE at doses up to 25 mg/kg/day via gavage for 70 days priorto mating through PND 21 (~21 weeks).Exposure to decaBDE at 1,800 mg/kg/day for 28 days did not cause increased pulmonary viral titers ofRSV (measured 5 days post-infection) in mice. In rat dams exposed to 300 mg/kg/day from 21 days priorto mating through PND 21, altered T-lymphocyte cell population distribution in the thymus and asignificantly reduced response to in vitro PHA exposure in cultured lymphocytes were observed. In another study, no dose-related changes were reported for T-cell, B-cell, or macrophage population PBDEs 242. RELEVANCE TO PUBLIC HEALTHdistribution in the spleen of rats exposed to decaBDE at doses up to 60 mg/kg/day via gavage for 28 days. In a high-dose study, female mice exposed to decaBDE at 800 mg/kg every other day showed impaired CD4 T-cell function from 4 to 10 months of exposure, compared with controls. Chronic ingestion of decaBDE caused splenic lesions (hematopoiesis, fibrosis, lymphoid hyperplasia) in rats exposed to ≥1,200 mg/kg/day for 103 weeks. After exposure for 13 weeks, histopathologicalexaminations of spleen, thymus, lymph node, and/or bone marrow tissues showed no effects in rats ormice exposed to decaBDE at doses up 8,000 or 9,500 mg/kg/day. In rat dams exposed to decaBDE at 300 mg/kg/day for 21 days prior to mating through PND 21, lesions of the thymus (thickened thymuscapsule, decreased lymphoid tissue in the cortex with adipose tissue replacement, increased medulla size,and obscured corticomedullary junction) and spleen (decreased size and number of lymphoid nodules, thinner lymphatic sheath around arteries, and fibrotic tissue with macrophages in the medulla) wereobserved.Evidence from animals suggests that PBDE exposure may cause immune suppression, but data are limited and inconsistent. Additionally, comprehensive immunological evaluations have not been performed and human data are extremely limited. Therefore, currently available information is insufficient to adequatelycharacterize the human immunotoxic potential of PBDEs.Cancer. In human case-control epidemiological studies, no clear associations have been found betweennon-Hodgkin’s lymphoma risk and exposure to BDE 47 in a group of Swedish men and women including19 cases and 27 controls, testicular cancer risk and serum PBDE (sum of BDE 47, BDE 99, and BDE 153) in a small group of Swedish men and women including 58 cases and 58 controls, breast cancerrisk and adipose concentrations of PBDE (BDE 47, BDE 99, BDE 100, BDE 153, BDE 154, and theirsum) in a group of women from the San Francisco Bay area of California including 78 cases and 56 controls, breast cancer and BDE 47 in serum in native Alaskan women, thyroid cancer and serumPBDE (BDE 47, BDE 99, BDE 100, BDE 153, and their sum) in participants in a large multicenterclinical trial in the United States that included 104 cases and 208 controls, or prostate cancer and serumlevels of BDE 47 in a study involving 208 prostate cancer incident cases and 268 controls in Singaporean males. In a study examining the association between exocrine pancreatic cancer risk and PBDE concentrations in adipose tissue (sum of BDE 28, BDE 47, BDE 66, BDE 100, BDE 99, BDE 85, BDE 154, BDE 153, BDE 138, and BDE 183) in a group of Swedish men and women, PBDEconcentrations were significantly higher in the 21 cases compared with the 59 controls. Case-controlanalysis found that the risk of pancreatic cancer was not significantly increased with lipid PBDE using PBDEs 252. RELEVANCE TO PUBLIC HEALTHmedian concentration in controls as a cut-off after adjustment for age, sex, and BMI at tissue sampling;however, the increase in risk was significant when the BMI adjustment was performed for the year before tissue sampling (body weight 1 year before tissue sampling obtained by questionnaire).For most PBDEs, including pentaBDE and octaBDE, animal studies of carcinogenic effects are notavailable; cancer data on PBDEs in animals are limited to results of studies on commercial decaBDEproducts. In a bioassay conducted by the National Toxicology Program (NTP), male and female rats wereexposed to high purity commercial decaBDE (lots that were 96 or 94–97% pure) in the diet in low dosesof 1,120 and 1,200 mg/kg/day, respectively, and high doses of 2,240 and 2,550 mg/kg/day, respectively,for 103 weeks. Male and female mice were similarly exposed to low doses of 3,200 and 3,760 mg/kg/day, respectively, and high doses of 6,650 and 7,780 mg/kg/day, respectively. Incidences ofneoplastic nodules in the liver were significantly increased in the male and female rats, although the termneoplastic nodule is poorly defined and understood, and is no longer used by NTP to characterizehepatoproliferative lesions in rats. Incidences of hepatocellular adenoma or carcinoma (combined) weresignificantly increased in the male mice. Slightly elevated incidences of thyroid gland follicular celladenoma or carcinoma (combined) were additionally observed in exposed male mice, although theincreases were not statistically significant. Carcinogenicity was also evaluated in rats that were exposedto 0.01, 0.1, or 1.0 mg/kg/day dietary doses of a 77.4% decaBDE mixture (containing 21.8% nonaBDEand 0.8% octaBDE) for approximately 2 years. No exposure-related neoplastic changes were found, butthe power of this study to detect carcinogenic effects is limited by the very low dose levels in comparisonto those tested in the NTP bioassay.The EPA hazard descriptor for decaBDE is “suggestive evidence of carcinogenic potential” based on:(1) no studies of cancer in humans exposed to decaBDE; (2) a statistically significant increase inincidence of neoplastic nodules and a slight increase in incidence of carcinomas (not statisticallysignificant) in the liver of low- and high-dose male rats and high-dose female rats; (3) a significantlyincreased incidence of hepatocellular adenoma or carcinoma (combined) in male mice at the low dose and marginally increased incidence at the high dose; (4) a nonsignificantly increased incidence ofhepatocellular adenoma or carcinoma (combined) in female mice; (5) a slightly greater (but statisticallynot significant) incidence of thyroid gland adenomas or carcinomas (combined) in dosed male and female mice; (6) a significantly increased incidence in male mice, at both doses, of follicular cell hyperplasia, considered by many as a precursor to thyroid tumors; and (7) an apparent absence of genotoxic potential. DecaBDE has been classified as a Group 3 carcinogen (not classifiable as to its carcinogenicity tohumans) by the International Agency for Research on Cancer (IARC) based on inadequate evidence of PBDEs 262. RELEVANCE TO PUBLIC HEALTHcarcinogenicity in humans and inadequate or limited evidence in experimental animals. The EPA assigns the cancer category Group D (not classifiable as to human carcinogenicity) to mono-, di-, tri-, tetra-, penta-, hexa-, octa-, and nonaBDEs and reports “inadequate information” to classify the specific congeners 2,2’,4,4’-tetraBDE, 2,2’,4,4’,5-pentaBDE, and 2,2’,4,4’,5,5’-hexaBDE. The Department ofHealth and Human Services has not evaluated PBDEs for carcinogenicity. ACGIH has no data regardingcancer classifications for PBDEs.2.3 MINIMAL RISK LEVELS (MRLs)Estimates of exposure levels posing minimal risk to humans (MRLs) have been made for PBDEs. AnMRL is defined as an estimate of daily human exposure to a substance that is likely to be without an appreciable risk of adverse effects (noncarcinogenic) over a specified duration of exposure. MRLs are derived when reliable and sufficient data exist to identify the target organ(s) of effect or the most sensitive health effect(s) for a specific duration within a given route of exposure. MRLs are based onnoncancerous health effects only and do not consider carcinogenic effects. MRLs can be derived foracute, intermediate, and chronic duration exposures for inhalation and oral routes. Appropriatemethodology does not exist to develop MRLs for dermal exposure.Although methods have been established to derive these levels (Barnes and Dourson 1988; EPA 1990), uncertainties are associated with these techniques. Furthermore, ATSDR acknowledges additionaluncertainties inherent in the application of the procedures to derive less than lifetime MRLs. As anexample, acute inhalation MRLs may not be protective for health effects that are delayed in developmentor are acquired following repeated acute insults, such as hypersensitivity reactions, asthma, or chronic bronchitis. As these kinds of health effects data become available and methods to assess levels ofsignificant human exposure improve, these MRLs will be revised.People are environmentally exposed to PBDE mixtures of different congeneric composition than previously used commercial PBDE products. Although the toxicity or potency of environmental mixturesof congeners consequently may be greater or less than that of the commercial PBDE mixtures orindividual congeners, there are insufficient mixture toxicity data on which to directly base MRLs forenvironmental PBDEs. Due to the likelihoods that (1) multiple mechanisms (aryl hydrocarbon receptor[AhR]-receptor-dependent mechanisms, AhR-independent mechanisms, or both) may be involved in health effects induced by PBDEs, (2) different PBDE congeners may produce effects by differentmechanisms, and (3) humans are exposed to complex mixtures of interacting PBDEs with differing PBDEs 272. RELEVANCE TO PUBLIC HEALTHbiological activities, as well as to the lack of a suitable approach for quantitatively evaluating joint toxic action from concurrent exposures to PBDEs, PBBs, PCBs, chlorinated dibenzo-p-dioxins (CDDs), and/orchlorinated dibenzofurans (CDFs) in the environment, data from previously used commercial PBDEmixtures and individual congeners were reviewed to develop MRLs for assessing health risks fromenvironmental exposures to PBDEs.Separate MRLs were derived for lower-brominated PBDEs and decaBDE based on important differences in pharmacokinetics and toxicity of decaBDE compared to lower-brominated PBDEs. The most recentand best available estimates of oral absorption efficiencies for PBDE congeners indicate a range of 10– 26% for decaBDE (BDE 209) and 70–85% for tetraBDE (BDE 47), pentaBDE (BDE 99, BDE 100), and hexaBDE (BDE 153, BDE 154) (Chen et al. 2006; Hakk et al. 2002a, 2002b, 2009; Morck and Klasson Wehler 2001; Morck et al. 2003; Örn and Klasson-Wehler 1998; Riu et al. 2008; Sandholm et al. 2003;Sanders et al. 2006a, 2006b; Staskal et al. 2005). Consistent with the higher absorption efficiencies of thelower-brominated congers, the animal toxicity database indicates that toxic effects occur at lower doses following exposure to lower-brominated PBDEs than following exposure to decaBDE. For example, theacute exposure levels required to cause neurobehavioral effects in mice are higher for decaBDE(≥2.22 mg/kg) than penta-, tetra-, and hexaBDE (≥0.8, ≥1, and ≥0.45 mg/kg, respectively) (Eriksson et al. 2001; Gee and Moser 2008; Johansson et al. 2008; Sand et al. 2004; Viberg et al. 2003a, 2003b, 2004a, 2004b). Inhalation MRLsLower-brominated BDEs. Derivation of an acute-duration MRL for lower-brominated BDEs is notrecommended at this time due to insufficient information. The inhalation database for acute-duration exposure to PBDEs is essentially limited to a single 14-day unpublished industry-sponsored study ofoctaBDE in rats (Great Lakes Chemical Corporation 1978). In this study, groups of five male and fivefemale Charles River CD rats were whole-body exposed to dust of an unspecified commercial octaBDEmixture in mean analytical concentrations of 0, 0.6, 3.7, 23.9, or 165.2 mg/m3 for 8 hours/day for14 consecutive days. The average mass median aerodynamic diameter (MMAD) and geometric standarddeviation (GSD) of the particles were 3.5 µm and 2, respectively. Study end points included clinicalsigns (including observations for respiratory distress and nasal and ocular irritation), body weight and food consumption, hematology (5 indices), blood chemistry (5 indices, thyroid hormones not assessed), urinalysis (10 indices), organ weights (5 organs including thyroid/parathyroid), gross pathology, and histology (21 tissues including nasal turbinates, trachea, lungs, and thyroid). The clinical laboratory tests PBDEs 282. RELEVANCE TO PUBLIC HEALTHwere limited to rats in the control and two highest dose groups. The histological exams were limited tothe control and highest dose groups, except for the liver, which was examined in all groups. Signs ofincreased respiration rate (rapid breathing) were observed by the end of each exposure period in rats exposed to ≥24 mg/m3; this effect always disappeared by the following morning. Liver weight wassignificantly increased and hepatic lesions occurred in rats exposed to ≥3.7 mg/m3. At 3.7 mg/m3, theliver lesions consisted of very slight to slight, focal to multifocal cytoplasmic enlargement of thehepatocytes, accompanied by focal acidophilic degeneration of individual to small groups of cells. The liver lesions were similar in the higher dose groups except that the hepatocyte enlargement was multifocalto diffuse in distribution, and there were focal, small to large areas of hepatocellular necrosis present to avery slight to marked degree. There were no exposure-related histological changes in other tissues.As detailed above, hepatocellular hypertrophy accompanied by some degenerative hepatocellular changes was found following exposure to octaBDE at concentrations ≥3.7 mg/m3 for 14 days (Great Lakes Chemical Corporation 1978). However, this study is limited by small animal numbers and incompleteevaluation of other end points at lower doses. Additionally, a well-designed 13-week study (Great Lakes Chemical Corporation 2000) found hepatocellular hypertrophy at a higher minimum effect level(16 mg/m3) than the 14-day study, but no degenerative liver changes. The available information indicates that there is insufficient evidence for considering the hepatic changes as adverse acute effects. More importantly, exposure to ≥16 mg/m3 caused changes in serum levels of thyroid hormones (decreased T3, increased TSH) in the 13-week study. Thyroid hormone levels were not determined in the 14-day study. Therefore, due to the lack thyroid hormone data in the 14-day study, as well as the lack of any clearlowest-observed-adverse-effect levels (LOAELs) for the other end points in the 14-day study, particularlyat exposures levels below the LOAEL for thyroid effects in the 13-week study, the data are inadequate to derive an MRL for acute-duration exposure.• An MRL of 0.006 mg/m3 has been derived for intermediate-duration inhalation exposure (15– 364 days) to lower-brominated BDEs.The intermediate-duration inhalation MRL is based on a no-observed-adverse-effect level (NOAEL) of 1.1 mg/m3 for changes in thyroid hormones in rats that were intermittently exposed to octaBDE for13 weeks (Great Lakes Chemical Corporation 2000). Calculation of the MRL is detailed below.The inhalation database for intermediate-duration exposure to PBDEs consists of one well-conducted13-week study (Great Lakes Chemical Corporation 2000). This is an unpublished industry-sponsored study in which a commercial octaBDE product (bromine content 78.7%) was administered to groups of PBDEs 292. RELEVANCE TO PUBLIC HEALTH10 male and 10 female Crl:CD(SD)IGS BR rats, via nose-only inhalation as a dust aerosol, in measured concentrations of 0 (air only), 1.1, 16, or 202 mg/m3 for 6 hours/day, 5 days/week, for 13 weeks. The mean MMADs in the low to high exposure groups were 2.0, 2.7, and 2.8 μm, and the corresponding mean GSDs were 3.37, 3.72, and 3.01. Clinical and physical signs, body weight, food consumption, and survival were evaluated throughout the study. Ophthalmic, hematology (11 indices), serum chemistry(18 indices), and serum thyroid hormone (TSH, total T3, and total T4) evaluations were performed nearthe end of the exposure period. Urine analyses were not conducted. Comprehensive necropsies, organ weight measurements, and histological examinations (including respiratory tract and thyroid) wereperformed following exposure termination. Hepatic, nasal, lung, thyroid, and ovarian effects were observed (Great Lakes Chemical Corporation2000). The liver was affected in both sexes as shown by dose-related increases in centrilobularhepatocellular hypertrophy at ≥16 mg/m3 and liver weight (absolute and relative) at 202 mg/m3. Totalincidences of centrilobular hepatocellular hypertrophy in the 0, 1.1, 16, and 202 mg/m3 groups were 1/10, 0/10, 3/10, and 10/10, respectively, in males and 0/10, 0/10, 3/10, and 6/10, respectively, in females;severity was predominantly minimal in affected animals from all groups. The incidence of nasal goblet cell lesions was increased at 202 mg/m3, but showed no clear dose-related trends for increasing incidence or severity. Total incidences of nasal goblet cell hypertrophy were slightly increased in nasal level II ofboth sexes at ≥1.1 mg/m3; respective incidences in the 0, 1.1, 16, and 202 mg/m3 exposure groups were4/10 (all minimal), 9/10 (7 minimal, 2 mild), 6/10 (all minimal), and 10/10 (9 minimal, 1 mild) in males, and 2/10 (all minimal), 6/10 (all minimal), 4/10 (all minimal), and 8/10 (all minimal) in females. Nasal goblet cell hypertrophy was also slightly increased in nasal level IV in males at 202 mg/m3 (4/10, 0/10, 1/10, and 8/10, all minimal severity, not increased in females). Histological changes in the lungs included alveolar histiocytosis and chronic active inflammation that were only clearly increased in incidence at202 mg/m3. Total incidences of alveolar histiocytosis at 0, 1.1, 16, and 202 mg/m3 were 3/10, 5/10, 5/10, and 10/10, respectively, in males, and 0/10, 5/10, 2/10, and 10/10, respectively, in females.Corresponding total incidences of chronic active lung inflammation were 0/10, 0/10, 2/10, and 10/10 in males, and 0/10, 1/10, 1/10, and 10/10 in females. The severity of both lesion types tended to increase from minimal at lower doses to mild/moderate at 202 mg/m3. Gross lung changes also occurred in both sexes at 202 mg/m3; these included lung firmness and white discoloration and/or enlargement in thebronchial and/or mediastinal lymph nodes. The lymph node effects correlated with the histologicalfinding of granulomatous inflammation. There were no exposure-related gross or histopathological changes in the spleen, bone marrow, thymus, or other tissues, including thyroid. Thyroid hormoneassessments, however, showed exposure-related decreases in mean thyroxine (total T4) at ≥16 mg/m3 in PBDEs 302. RELEVANCE TO PUBLIC HEALTHboth sexes, and increases in TSH at ≥16 mg/m3 in males and 202 mg/m3 in females. The changes were usually statistically significant (p<0.05 or p<0.01) compared to controls and were considered to beconsistent with chemical-induced hypothyroidism. There were no serum T3 changes. Qualitative histological evaluations of step sections of ovaries showed an absence of corpora lutea in 3/10 females exposed to 202 mg/m3, compared to 0/10 in the control and lower exposure groups. This 30% incidencewas interpreted to be a treatment-related effect because an absence of corpora lutea was consideredunusual in rats at 20 weeks of age.Considering the minimal severity of the nasal goblet cell hypertrophy, lack of clear dose-relatedincreasing trends for incidences and severity of this nasal effect, clear identification of both a NOAEL(1.1 mg/m3) and LOAEL (16 mg/m3) for changes in serum levels of thyroid hormones, and abundantevidence for thyroid effects of PBDEs in oral studies, the effects on thyroid hormones are the mostappropriate basis for estimation of an intermediate-duration inhalation MRL. The MRL of 0.006 mg/m3 was derived by dividing the NOAELHEC of 0.53 mg/m3 by an uncertainty factor of 30 (3 for species to species extrapolation with dosimetric adjustments and 10 for human variability) and a modifying factor of3 (for an incomplete database reflecting a single study in one species). The NOAELHEC was calculatedusing the following equations:NOAELADJ = 1.1 mg/m3 x 6 hours/24 hours x 5 days/7 days = 0.196 mg/m3 NOAELHEC = NOAELADJ x RDDR = 0.196 mg/m3 x 2.7 = 0.53 mg/m3The regional deposited dose ratio (RDDR) for the extrathoracic region was used to extrapolate deposited doses in rats to deposited doses in humans. The following parameters were used to calculate the RDDR:MMAD of 2.0 μm with a mean GSD (sigma g) of 3.37, default human body weight of 70 kg, and adefault female F344 rat body weight of 0.18 kg. Additional information on the derivation of theintermediate-duration inhalation MRL for lower-brominated BDEs is provided in Appendix A.No MRL was derived for chronic-duration inhalation exposure to lower-brominated BDEs due to a lackof chronic studies.Decabromodiphenyl Ether. No MRLs were derived for acute-, intermediate-, or chronic-duration inhalation exposure to decaBDE due to a lack of inhalation studies on this PBDE congener. PBDEs 312. RELEVANCE TO PUBLIC HEALTHOral MRLsLower-brominated Diphenyl Ethers• An MRL of 0.00006 mg/kg/day has been derived for acute-duration oral exposure (14 days orless) to lower-brominated diphenyl ethers.The acute oral MRL is based on a LOAEL of 0.06 mg/kg/day for endocrine effects in rat dams andreproductive and neurobehavioral effects in F1 offspring exposed to 2,2’,4,4’,5-pentaBDE (BDE 99) on GD 6 via gavage (Kuriyama et al. 2005, 2007; Talsness et al. 2005). The MRL was estimated by dividingthe 0.06 mg/kg LOAEL by an uncertainty factor of 1,000 (10 for use of a LOAEL, 10 for animal to human extrapolation, and 10 for human variability). In a series of reports, pregnant rats were exposed to BDE 99 at 0, 0.06, or 0.3 mg/kg via gavage on GD 6 (Kuriyama et al. 2005, 2007; Talsness et al. 2005). Serum thyroid hormones levels (T3, free-T3, T4, freeT4) were analyzed in dams and pups on PNDs 1, 14, and 22 (Kuriyama et al. 2007). Male and femaleoffspring were evaluated for emergence of physical landmarks and reflexes and for open-field behavioron PNDs 36 and 71 (Kuriyama et al. 2005). Twelve male offspring per dose group were sacrificed atPND 140, and the thymus, spleen, liver, testis, epididymis, seminal vesicle, and ventral prostate were weighed. The right testis and caudal epididymis were retained for spermatid and sperm counts andmorphology, respectively. Additionally, blood was collected for analysis of testosterone and LH levels(Kuriyama et al. 2005). Similarly, 10 F1 females per group were sacrificed ~PND 90 for histologicalevaluation of the ovary, uterus, and vagina. Ovarian follicles were counted in 10 ovaries from each group, and 1 ovary from 1 female offspring in each group was analyzed by transmission electronmicroscopy (Talsness et al. 2005). Fertility was assessed in F1 males and females (20/group) mated tounexposed partners. The uterine and F2 fetal weights and the number of implantations, resorptions, and fetuses were determined. The F2 fetuses were examined for external anomalies and when present, thefetuses were stained and examined for skeletal anomalies (Kuriyama et al. 2005; Talsness et al. 2005). Ina separate group of F1 males, male sexual behavior was assessed in 20 males/group at ~PND 160 (Kuriyama et al. 2005). Serum T4 levels were significantly decreased by 23–33% in the 0.06 and 0.3 mg/kg dams, sacrificed on PND 1. No changes were observed in T3, free-T3, or free-T4 at PND 1 or any thyroid hormone levels atPND 22 in dams. In pups, no dose-related changes were observed at PND 1 or 14. At PND 22, serum T4 was significantly decreased by 19–22% in F1 males and females and serum free-T4 was significantly PBDEs 322. RELEVANCE TO PUBLIC HEALTHdecreased by 24% in F1 females exposed to 0.3 mg/kg (Kuriyama et al. 2007). For F1 development ofphysical landmarks and reflexes, no exposure-related effects were observed for the age at furdevelopment or eye opening, testes descent, or the ability to master the rotating rod test. However, significant delays in the eruption of incisors in F1 pups and the development of the cliff-drop aversion reflex were observed in F1 males in the 0.3 mg/kg group, compared with controls. During a 24-hourobservation of open-field activity, total activity, time spent active, duration of activity per active phase,and total activity per active phase were all significantly increased in F1 offspring on PND 36 in the0.3 mg/kg group, compared with controls. On PND 71, the increased total activity and time spent activepersisted in the 0.3 mg/kg group, and was also significantly increased in the 0.06 mg/kg group. In F1 males sacrificed on ~PND 140, no exposure-related changes were observed in body weight, liverweight, or thymus weight; however, absolute spleen weight was significantly increased by 9% in the 0.06 and 0.3 mg/kg groups, and relative spleen weight was significantly increased by 12% in the0.06 mg/kg group. Compared with controls, significantly altered male reproductive organ weights atPND 140 included a 10 and 11% decrease in relative testes and epididymis weight, respectively, in the0.3 mg/kg group and a 5% decrease in relative epididymis weight in the 0.06 mg/kg group; no significantchanges were observed in absolute organ weights. In both dose groups, the number of spermatids and sperm and daily sperm production were significantly decreased, compared with controls. No exposure-related effects were observed for sperm morphology. No changes were observed in serum testosterone orLH levels. Despite sperm alterations, no significant exposure-related effects were observed in male reproductive function or the majority of male sexual behaviors. The only significantly altered male sexual behavior was a 32% decrease in the percent of males with two or more ejaculations. In F1 females sacrificed on ~PND 90, no statistically significant, exposure-related histological changes were observed at the light microscopic level in the ovary, uterus, or vagina of female offspring, and no exposure-related effects were observed in the number of ovarian follicles. However, multipleultrastructural changes were noted in the ovaries of PND 90 female offspring from dams exposed to 0.06 or 0.3 mg/kg, including destruction of the surface of the serosal epithelial cells, necrosis, andnumerous vesicular structures with dense granular material within the cytoplasm. Additional changes observed in the 0.3 mg/kg group included degenerative changes and aggregates of small and large vesicles filled with homogeneously dense granular material in the cytoplasm and clumped chromatinwithin the condensed nucleus. No exposure-related changes were found for F1 female pregnancy rate,total implantation sites, implantation sites/dam, F2 fetuses/gravid dam, or total number of live F2 fetuses. However, the resorption rates were 12 and 15% in the 0.06 and 0.03 mg/kg groups, respectively, PBDEs 332. RELEVANCE TO PUBLIC HEALTHcompared with the control rate of 9%. Statistics were not reported; however, the resorption rates in the exposed rats were also reportedly increased compared with historical controls (average control resorptionrate=5.4%, with rates up to 10% considered to be within normal limits). In addition, the percentage oflitters with resorptions was higher in the exposed females, being 47% in the control group and 69% and 72% in the 0.06 and 0.3 mg/kg groups, respectively. In F2 pups, mean fetal weight was significantlyincreased (by 5%) in the 0.06 mg/kg group, but not in the 0.3 mg/kg group, compared with controls. Three fetuses from different litters in the 0.3 mg/kg/day group showed skeletal anomalies (tail, skull, vertebrae); however, this incidence of anomalies in 3/18 litters is not significantly elevated compared withthe control incidence of 0/19 (Fisher’s exact test, performed for this review). Collectively, these studies indicate a LOAEL of 0.06 mg/kg, the lowest dose tested, for endocrine effectsin F0 dams (decreased serum T4), reproductive effects in F1 adult offspring (impaired spermatogenesis, ultrastructural changes in ovaries, increased resorptions in F1 females mated to unexposed males) andneurobehavioral effects in F1 adult offspring (increased activity in open field). No NOAEL wasidentified. Data from other several acute-duration studies of PBDEs support the selection of the co-critical effects observed at the LOAEL of 0.06 mg/kg:• Numerous studies reported reduced serum T4 levels in adult, nonpregnant mice and rats followingacute exposure to commercial pentaBDE mixtures (Bromkal 70, Bromkal 70-5 DE, DE-71), thecommercial octaBDE mixture DE-79, or 2,2’,4,4’-tetraBDE (BDE 47). Significant reductions of19–92% have been reported following gavage exposure at doses ≥10 and ≥0.8 mg/kg/day in ratsand mice, respectively, for 1–14 days (Darnerud and Sinjari 1996; Fowles et al. 1994; Hallgren and Darnerud 1998, 2002; Hallgren et al. 2001; Hoppe and Carey 2007; Richardson et al. 2008;Stoker et al. 2004, 2005; Zhou et al. 2001).• In a companion study to the critical studies described above, pregnant rats (8/group) wereadministered BDE 47 (98% purity) at 0, 0.14, or 0.7 mg/kg via gavage in peanut oil vehicle on GD 6 (Talsness et al. 2008). As observed in pentaBDE-exposed F1 females, ultrastructuralchanges (accumulation of vesicular structures with homogeneously dense granular material in the cytoplasm of the stromal cells, large vacuoles) were observed in the ovaries of F1 females fromboth dose groups on PND 100. No exposure-related changes were observed in F1 female fertilityor F2 litter parameters. F1 males were not evaluated for developmental reproductive effects following tetraBDE exposure.• Alterations in open-field activity have been consistently reported in mice exposed to pentaBDE(BDE 99) at doses ≥0.8 mg/kg on PND 3 or 10 and evaluated at 2–8 months of age, characterized by decreased activity during the first 20-minute period of a 1-hour session, followed by increased activity during the third 20-minute period (Eriksson et al. 2002b, 2006; Fischer et al. 2008; Sand et al. 2004; Viberg et al. 2002, 2004a, 2004b). Several other 1-day exposure studies reported PBDEs 342. RELEVANCE TO PUBLIC HEALTHsimilar findings in rats and mice following exposure to various lower-brominated PBDEs. Decreased spontaneous activity and/or impaired habituation were observed in rats exposed to BDE 99 at 8 mg/kg on PND 10, mice exposed to 2’,4,4’,5,5’-hexaBDE (BDE 153) at≥0.45 mg/kg on PND 10, mice exposed to BDE 47 at 10.5 mg/kg on PND 10, mice exposed to 2,2’,3,4,4’,5’,6-heptaBDE (BDE 183) at 15.2 mg/kg on PND 3, and mice exposed 2,2’,3,4,4’,5,5’,6-octaBDE (BDE 203) at 16.8 mg/kg on PND 3 or 10 (Eriksson et al. 2001; Viberg et al. 2003a, 2005, 2006). Increased vertical activity was significantly increased at4 months, but not at 2 months, in mice exposed to BDE 47 at ≥1 mg/kg on PND 10; no changeswere observed in horizontal activity or habituation (Gee and Moser 2008). Additional information on the derivation of the acute-duration oral MRL for lower-brominated BDEs is provided in Appendix A.• An MRL of 0.000003 mg/kg/day has been derived for intermediate-duration oral exposure (15– 364 days) to lower-brominated BDEs.The intermediate oral MRL is based on a minimal LOAEL of 0.001 mg/kg/day for a 34% reduction inserum testosterone in male rats exposed to 2,2’,4,4’-tetraBDE (BDE 47) for 8 weeks via gavage (Zhang etal. 2013b). The MRL was estimated by dividing the 0.001 mg/kg/day minimal LOAEL by an uncertaintyfactor of 300 (3 for use of a minimal LOAEL, 10 for animal to human extrapolation, and 10 for human variability).Groups of 20 male rats were exposed to BDE 47 (≥98.7%) at 0, 0.001, 0.03, or 1 mg/kg/day via gavage in corn oil 6 days/week for 8 weeks (Zhang et al. 2013b). Twenty-four hours after the final treatment, rats were sacrificed. Testes were fixed for histological analysis and labeling of apoptotic cells or prepared foranalysis of sperm production. Daily sperm production was estimated by dividing the total number ofmature spermatids per testis by 6.1 (i.e., the days of the seminiferous cycle that the spermatids are presentin the seminiferous epithelium). Testicular samples were examined for reactive oxygen species (ROS)and mRNA expression of apoptosis related proteins (ser15, ser473, p53, PTEN, AKT, BAD, caspase 3, FAS, FASL). Serum levels of E2, FSH, LH, and testosterone were measured. Histological examination of the testes showed a significant increase in the number of multinucleated giant cells (arising fromspermatocytes that aborted meiosis) at ≥0.03 mg/kg/day and abundant vacuolar spaces in the seminiferous epithelium at 1 mg/kg/day (quantitative data not reported). Additionally, the number of apoptotic cellswas significantly increased by 1.9- and 3-fold in the testes of rats from the 0.03 and 1 mg/kg/day groups, respectively, and the mRNA levels of several apoptosis genes were elevated in a dose-related manner.Daily sperm production was significantly decreased by 23% in the 1 mg/kg/day group, compared with controls. Serum testosterone was significantly decreased by ~34, 53, and 62% in the 0.001, 0.03, and1 mg/kg/day groups, respectively, compared with controls. No exposure-related changes were observed PBDEs 352. RELEVANCE TO PUBLIC HEALTHin serum E2, FSH, or LH levels. Testicular ROS levels were significantly elevated at 1 mg/kg/day,compared with controls. A minimal LOAEL of 0.001 mg/kg/day was determined for this study based on the 34% decrease in serum testosterone. The change in testosterone is considered a minimal LOAELbecause it is unclear if the magnitude of change represents a biologically adverse effect; however, this statistically significant reduction in serum testosterone is considered an early indication of damage to the male reproductive system, considering the additional effects observed at ≥0.03 mg/kg/day (histologicallesions in testes, sperm effects).One additional rat study and a mouse study report histopathological changes in the testes followingintermediate-duration exposure to tetraBDE at ≥0.03 mg/kg/day; neither study evaluated serumtestosterone levels (Huang et al. 2015; Wang et al. 2013). In the rat study, a NOAEL of 0.001 mg/kg/dayand a LOAEL of 0.03 mg/kg/day were identified for increased epithelial thickness and spermatocyte apoptosis in the testes of males exposed to BDE 47 for 8 weeks via gavage (Huang et al. 2015). In the mouse study, a NOAEL of 0.0015 mg/kg/day and a LOAEL of 0.045 mg/kg/day were identified for germcell loss and increased apoptosis in the testes of males exposed to BDE 47 for 30 days via gavage (Wang et al. 2013). No other study evaluated testicular histopathology following exposure to BDE 47. Followingintermediate exposure to other congeners, no changes in testicular histology were observed in rats exposed to commercial pentaBDE mixtures (Bromkal 70-5 DE, DE-71) at gavage doses up to 250 mg/kg/day for 15–28 days (Becker et al. 2012; Oberg et al. 2010), commercial penta- or octaBDEmixtures (DE-71, unspecified octa mixture) at dietary doses up to 750 mg/kg/day for 28–90 days (IRDC1976, 1977; WIL Research Laboratories 1984), or a pentaBDE mixture (52.1% pentBDE, 44.2%decaBDE, 0.4% octaBDE) at dietary doses up to 20 mg/kg/day for 70 days (Ernest et al. 2012). However,testicular apoptosis was not evaluated in any of these studies. No other study evaluated serum testosterone levels following exposure to BDE 47. However, as observed with exposure to BDE 47, acute exposure to 0.06 or 1.2 mg/kg of 2,2’,4,4’,5-pentaBDE (BDE 99) alsoled to a significant 40–45% decrease in serum testosterone levels in rats (Alonso et al. 2010). No otherstudies evaluated this end point following exposure to single congeners. Other studies evaluating serumtestosterone levels after intermediate-duration exposure to lower-brominated PBDEs mixtures (DE-71, dietary PBDE mixture described above) did not report exposure-related decreases (Becker et al. 2012;Ernest et al. 2012; Stoker et al. 2005). These data suggest that the individual congeners, BDE 47 andBDE 99, which have been identified as two of the most abundant congeners for human exposure (Harrad PBDEs 362. RELEVANCE TO PUBLIC HEALTHet al. 2004; Lorber 2008; Wong et al. 2013), may have a greater capacity to alter serum testosterone levels than PBDE mixtures.One-generation studies of BDE 47 reported developmental effects at ≥0.03 mg/kg/day, including:• Impaired spatial learning in the Barnes maze in PNW 8 offspring of mouse dams fed tetraBDEdosed cornflakes from pre-mating day 28 through PND 21 (Koenig et al. 2012).• Decreased center-field activity in an open field (indicating increased anxiety) in PND 60 femaleoffspring from mouse dams fed tetraBDE-dosed cornflakes from pre-mating day 28 through PND 21 (Ta et al. 2011).• Decreased pre-weaning weight, decreased pup vocalizations on PNDs 8–10, and decreased sociability on PND 72 in female offspring of mouse dams exposed to tetraBDE via gavage frompre-mating day 28 through PND 21 (Woods et al. 2012).Additional information on the derivation of the intermediate-duration oral MRL for PBDEs is provided inAppendix A.A chronic-duration oral MRL was not derived for lower-brominated PBDEs due to insufficient data. Only one chronic study of PBDEs other than high-purity decaBDE has been conducted (Kociba et al. 1975; Norris et al. 1975a). In this study, Sprague-Dawley rats (25/sex/dose level) were fed a 77.4% pure commercial decaBDE mixture (containing 21.8% nonaBDE and 0.8% octaBDE) for approximately2 years. Evaluations that included clinical signs, body weight, food consumption, hematology, clinicalchemistry, urine indices, and comprehensive histological examinations showed no exposure-relatedeffects. The highest NOAEL is 1 mg/kg/day (highest tested dose), but this NOAEL is not appropriate forMRL estimation due to insufficient sensitivity of the study. In particular, using the NOAEL of1 mg/kg/day and an uncertainty factor of 100, a chronic oral MRL based on this study would be 5 times higher than the 0.002 mg/kg/day intermediate MRL. A similar pattern was observed for thyroid effects in the study used to derive the acute-duration oral MRL (Zhou et al. 2001) as summarized above. Due to theinsufficiencies of the chronic data for MRL derivation, the intermediate oral MRL could be used as a value for chronic exposure.Decabromodiphenyl Ether• An MRL of 0.01 mg/kg/day has been derived for acute-duration oral exposure (14 days or less) to decabromodiphenyl ether PBDEs 372. RELEVANCE TO PUBLIC HEALTHThe MRL was derived based on a NOAEL of 1.34 mg/kg for neurobehavioral effects in 2–4-month-old mice following a single exposure to 2,2’,3,3’,4,4’,5,5’,6,6’-decaBDE (BDE 209) on PND 3 (Buratovic et al. 2014; Johansson et al. 2008). The MRL was estimated by dividing the 1.34 mg/kg NOAEL by an uncertainty factor of 100 (10 for animal to human extrapolation and 10 for human variability). In the first study (Johansson et al. 2008), neonatal male mice (3–4 litters/group) were exposed to a singledose of BDE 209 (98% purity) at 0, 1.34, 2.22, 13.4, or 20.1 mg/kg via gavage in a 20% fat emulsion vehicle (1:10 mixture egg lecithin and peanut oil) on PND 3 (Johansson et al. 2008). Mice were observed for clinical signs of toxicity and body weight was measured at PND 3 and PNW 4. Spontaneous motorbehavior (locomotion, rearing, total activity) was evaluated in an open field test at 2 months(10 mice/group) and at 4 months (16 mice/group). Motor activity was measured during a 60-minuteperiod, divided into three 20-minute intervals. Nicotine-induced behavior was evaluated at 4 monthsfollowing single subcutaneous injections of 80 µg nicotine/kg (8/group) or 10 mL 0.9% NaCl/kg(8/group). Anxiety was assessed at 4 months using the elevated plus maze. No clinical signs of toxicityor body weight effects were observed. At 2 months, significantly decreased locomotion, rearing, and totalactivity were observed during the first 20-minute interval of the open field assessment in mice exposed to≥2.22 mg/kg, compared with controls. However, during the third 20-minute interval, when activityshould decrease due to habituation, locomotion, rearing, and total activity were significantly increased inmice exposed to ≥13.4 mg/kg. None of the end points measured were significantly altered in mice exposed to 1.34 mg/kg. At 4 months, significantly decreased locomotion, rearing, and total activity wereobserved during the first interval of the open field assessment in mice exposed to ≥2.22 mg/kg, compared with controls. During the third interval, significantly increased locomotion, rearing, and total activitywere observed in mice exposed to ≥2.22 mg/kg. Additionally, total activity, but not rearing orlocomotion, was significantly decreased during the first 20-minute interval in the 1.34 mg/kg group; no significant changes were observed during the third interval in the 1.34 mg/kg group. Statistical analysisshows that habituation ability declined in mice exposed to ≥2.22 mg/kg/day when tested at 4 months ofage, compared with 2 months of age. At 4 months, nicotine exposure caused significantly decreasedactivity during the first interval in mice exposed to ≥13.4 mg/kg, compared with saline-injected mice from the same decaBDE exposure group. This finding is the opposite of the expected increase in activitydue to nicotine exposure, which was observed in controls and lower dose decaBDE groups. During thethird interval, mice exposed to ≥13.4 mg/kg and nicotine showed impaired habituation. No exposure-related effects were observed in the elevated plus maze assessment. A NOAEL of 1.34 mg/kg and aLOAEL of 2.22 mg/kg were determined for the nonhabituating profile (i.e., decreased activity early in thetest period and increased activity late in the test period). The singular finding of decreased total activity PBDEs 382. RELEVANCE TO PUBLIC HEALTHduring the first 20-minute interval at 4 months in the 1.34 mg/kg group was not considered sufficient to establish a LOAEL of 1.34 mg/kg. The nonhabituating profile, which is a common effect observed with developmental PBDE exposure (Eriksson et al. 2002b, 2006; Fischer et al. 2008; Sand et al. 2004; Viberget al. 2002, 2003a, 2004a, 2004b), was considered to be a stronger basis for a NOAEL/LOAELdetermination.In the second study (Buratovic et al. 2014), neonatal male mice (6 litters/group; 31–40 males and 23– 34 females per group) were administered 2,2’,3,3’,4,4’,5,5’,6,6’-decaBDE (BDE 209, >95% purity) atdoses of 0, 1.34, 5.76, or 13.4 mg/kg via gavage in a 20% fat emulsion vehicle (1:10 mixture egg lecithinand peanut oil) on PND 3. Mice were observed for clinical signs of toxicity and body weight changes throughout the study (no further details were provided). Spontaneous motor behavior (locomotion, rearing, total activity) was evaluated in an open field at 2 months (18/sex/group). Motor activity wasmeasured during a 60-minute period, divided into three 20-minute intervals. Directly after spontaneous motor evaluation, 9/sex/group were injected with a cholinergic agent (0.25 mg/kg paraoxon in males, 80 µg/kg nicotine in females), while the other 9/sex/group were injected with 0.9% saline, for evaluation of cholinergic-induced locomotion. At 4 months, spontaneous behavior was assessed again in the saline-injected animals only (9 males/group at all doses and 9 females/group in the control and high-dose group only). Learning and memory was assessed using the Morris water maze at 5 and 7 months in 13– 15 males from the 0, 5.76, and 13.4 mg/kg groups only (the same mice were evaluated at each time point). In the spontaneous activity assessment, a dose-related decrease in locomotion, rearing, and total activitywas observed during the first 20 minutes of open field testing in a novel environment at 2 months. Decreases were significant at all doses tested in both sexes; however, findings were only dose-related for total activity. However, during the third 20-minute interval, when activity should decrease due to habituation, locomotion, rearing, and total activity were significantly increased in males and females at ≥5.76 mg/kg. At 2 months, cholinergic agents caused decreased activity during the first interval in miceexposed to ≥5.76 mg/kg, compared with saline-injected mice from the same decaBDE exposure group.This finding is the opposite of the expected increase in activity due to paraoxon or nicotine exposure, which was observed in controls and low-dose decaBDE groups. During the third interval, mice exposed to ≥5.76 mg/kg and cholinergic agent showed impaired habituation. At 4 months, total activity during thefirst 20 minutes was still significantly decreased at all doses in males, and locomotion and rearing weresignificantly decreased in males in the mid- and high-dose groups only; all three parameters were significantly decreased in high-dose females (other doses not evaluated). All three parameters were significantly increased in high-dose males and females during the third 20-minute period, indicatingdecreased habituation; locomotion and rearing were also slightly, but significantly, increased in mid-dose PBDEs 392. RELEVANCE TO PUBLIC HEALTHmales. In the Morris water maze, initial learning was comparable between exposed and control mice at5 and 7 months. However, latencies to find the escape platform during the reversal learning phase(learning to find the escape platform in a new location after initial training) were significantly longer inmid- and high-dose males at 5 and 7 months (other exposure groups not assessed). A NOAEL of1.34 mg/kg and a LOAEL of 5.76 mg/kg were determined for the nonhabituating profile (i.e., decreased activity early in the test period and increased activity late in the test period). Similar to the Johansson etal. (2008) study, the finding of decreased total activity during the first 20-minute interval at 2 and 4 months in the 1.34 mg/kg group was not considered sufficient to establish a LOAEL of 1.34 mg/kg.The nonhabituating profile was considered to be a stronger basis for a NOAEL/LOAEL determination, and additional neurological effects (impaired learning, altered response to cholinergic agents) support a LOAEL of 5.76 mg/kg.A similar study supports the LOAEL of 2.22 mg/kg for altered neurobehavior in developmentallyexposed mice. Decreased spontaneous activity and impaired habituation were observed in 2- and 6-month-old mice exposed to BDE 209 at doses ≥2.22 mg/kg on PND 3, which was the lowest dose tested(Viberg et al. 2003b). These effects were not observed if exposure was on PND 10 or 19 at doses up to 20.1 mg/kg (Viberg et al. 2003b). Additionally, decreased spontaneous activity was observed in 2-month-old rats following exposure to BDE 209 doses ≥6.7 mg/kg on PND 3 (lowest dose tested)(Viberg et al. 2007). At 20.1 mg/kg, impaired habituation and decreased nicotine-induced behavior werealso observed. This nonhabituating behavior profile (i.e., decreased activity early in the test period and increased activity late in the test period) is consistent with neurobehavioral alterations observed followingearly postnatal exposure to lower-brominated PBDEs and has been reported in adult mice neonatallyexposed to certain PCB congeners (see the Acute MRL Worksheet for lower-brominated PBDEs for moredetails).Additional neurodevelopmental effects observed in mice following acute exposure to BDE 209 from PND 2 to 15 at 20 mg/kg/day via micropipette include delayed ontogeny of reflexes, increasedlocomotion in males at PND 70, and learning impairment and impulsivity at 16 months, but not at3 months (Rice et al. 2007, 2009). In rats, impaired learning was observed in Morris water maze inPND 25 rat offspring of dams exposed to BDE 209 from GD 1 to 14 at doses ≥30 mg/kg/day via gavage(Chen et al. 2014).Additional information on the derivation of the acute-duration oral MRL for BDE 209 is provided in Appendix A. PBDEs 402. RELEVANCE TO PUBLIC HEALTH• An MRL of 0.0002 mg/kg/day has been derived for intermediate-duration oral exposure (15– 364 days) to decabromodiphenyl ether.The MRL was derived based on a minimal LOAEL of 0.05 mg/kg/day for a 12% increase in serum glucose in adult rats exposed to 2,2′,3,3′,4,4′,5,5′,6,6′-decabromodiphenyl ether (BDE 209) for 8 weeksvia gavage (Zhang et al. 2013a). The MRL was estimated by dividing the 0.05 mg/kg/day LOAEL by anuncertainty factor of 300 (3 for use of a minimal LOAEL, 10 for animal to human extrapolation, and 10for human variability).Groups of 10 male rats were exposed to BDE 209 at 0, 0.05, 1, or 20 mg/kg/day daily via gavage in corn oil for 8 weeks (Zhang et al. 2013a). Rats were observed for clinical signs of toxicity and body weightswere measured every 3 days. Rats were fasted for 24 hours after the final gavage treatment, and then sacrificed. Body weights and heart, spleen, lung, kidney, and liver weights were recorded. Blood wascollected for clinical chemistry analysis (serum total cholesterol, triglycerides, glucose, insulin, and TNF-α) and determination of plasma markers of oxidative stress (malondialdehyde [MDE], reducedglutathione [GSH], and superoxidase dismutatase [SOD]). Liver samples from three rats in the controland low-dose (0.05 mg/kg/day) groups were collected for microarray analysis (Affymetrix GeneChip), and gene ontogeny category, pathway, gene-act-network, and gene co-expression analyses were conducted. Quantitative real-time-polymerase chain reaction (qPCR) was performed to quantitate geneexpression to validate the gene expression data obtained from microarray analysis. No clinical signs oftoxicity or body weight effects were observed. The relative liver weight was significantly decreased at1 and 20 mg/kg/day by 9% (absolute liver weights were not reported). No changes were observed in relative weights of heart, spleen, lung, or kidney. No exposure-related changes were reported in serumcholesterol or triglyceride levels. Serum glucose levels were significantly increased by 12, 18, and 21%in the 0.05, 1, and 20 mg/kg/day groups, compared with controls. Serum insulin was significantlydecreased by 50–60% at 1 and 20 mg/kg/day. Subsequent to this finding, the pancreas was evaluated histologically. Consistent with the insulin findings, morphological changes were seen at 1 and 20 mg/kg/day, including blurred boundaries among pancreatic islet cells (quantitative data not reported). Plasma SOD activity was significantly decreased in all exposed groups and plasma GSH was significantlydecreased at 1 and 20 mg/kg/day. Serum TNF-α was significantly increased at 1 and 20 mg/kg/day.Additionally, decaBDE induced 1,257 liver gene transcript changes, and 18 canonical pathways weresignificantly enriched. Four of them were involved in immune diseases, including autoimmune thyroid disease, graft-versus-host disease, allograft rejection, and T1DM. Subsequently, gene act network and gene coexpression network found that some major histocompatibility complex molecules and TNF-α PBDEs 412. RELEVANCE TO PUBLIC HEALTHwere involved in the T1DM pathway. A minimal LOAEL of 0.05 mg/kg/day was determined for thisstudy based on the 12% increase in serum glucose levels. The change in glucose is considered a minimalLOAEL because it is unclear if the magnitude of change represents a biologically adverse effect;however, the increase in serum glucose is considered to be part of a spectrum of effects indicative ofaltered insulin homeostasis and toxicity to the pancreas, including decreased serum insulin and morphological changes in pancreatic islet cells observed at ≥1 mg/kg/day following decaBDE exposure. The relevance of these findings to human health is uncertain. An analysis of cross-sectional NHANESdata showed a significant increase in risk of diabetes associated with serum concentrations of BDE 153 (but not other congeners), although the risk was higher with exposure to 50–75th percentile BDE 153concentrations than >75th percentile BDE 153 concentrations (Lim et al. 2008). Serum BDE 153 concentrations (but not BDE 28, BDE 47, BDE 85, BDE 99, BDE 100, or BDE 154) were also shown to be significantly associated with increased odds of developing gestational diabetes in a cohort of258 pregnant women (Smarr et al. 2016). However, other cross-sectional and prospective studies foundno relationship between serum PBDE concentrations and diabetes in an adult cohort from Wisconsin (Turyk et al. 2015), an elderly cohort in Finland (Airaksinen et al. 2011), or an elderly cohort in Sweden (Lee et al. 2011).Only one other animal study evaluated the pancreas following decaBDE exposure. In rats exposed to BDE 209 via gavage for 28 days at doses of 0, 0.27, 0.82, 2.47, 7.4, 22.2, 66.7, or 200 mg/kg/day, slightor moderate insulitis was observed in the Langerhan’s islets of the “majority of samples,” but findings were not exposure-related (Van der ven et al. 2008a). Similarly, no exposure-related effects were observed for serum glucose levels (Van der ven et al. 2008a). The only other study evaluating serumglucose levels after decaBDE exposure instead reported reduced serum glucose levels in male rats exposed to 20 mg/kg/day of a dietary PBDE mixture containing 52.1% pentaBDE (DE-71), 44.2%decaBDE (BDE 209), and 0.4% octaBDE (DE-79) for 70 days (Ernest et al. 2012). The observed decreased glucose levels could be due to the pentaBDE component, as male rats exposed to pentaBDE atdoses of 0.27–200 mg/kg/day for 28 days also showed decreased glucose levels; the study authors did notreport the lowest dose at which glucose levels were significantly lower in male rats, but they reported a BMD10RD of 179.55 mg/kg/day and a BMDL10RD of 66.7 mg/kg/day (Van der ven et al. 2008b). Othereffects occurred at doses 4–40-fold higher than the observed pancreatic and related effects:• A LOAEL of 2 mg/kg/day was identified for transient histopathological effects in the liver ofmale offspring and kidney of female offspring of rat dams exposed to BDE 209 from GD 10 toPND 21 (no NOAEL identified) (Fujimoto et al. 2011). PBDEs 422. RELEVANCE TO PUBLIC HEALTH• A LOAEL of 10 mg/kg/day was identified for hepatocytic swelling in the liver, vacuolization in the interstitial cells of testes, and sperm damage in PND 71 male offspring of mouse dams exposed to BDE 209 from GD 0 to 17 (no NOAEL identified) (Tseng et al. 2008, 2013).• A LOAEL of 20 mg/kg-day was identified for decreased anxiety in mice treated with BDE 209by daily gavage for 15 days (no NOAEL identified) (Heredia et al. 2012).• A LOAEL of 20.1 mg/kg/day was identified for altered hippocampal electrophysiology in ratsexposed to BDE 209 on GD 1 to PND 41, PNDs 1–21, or PNDs 22–41 (no NOAEL identified)(Xing et al. 2009).Additional information on the derivation of the intermediate-duration oral MRL for decaBDE is providedin Appendix A.No MRL was derived for chronic-duration oral exposure to decaBDE. Only one chronic study of high-purity decaBDE has been conducted. In this study, F344 rats and B6C3F1 mice (50/sex/group perspecies) were administered a commercial decaBDE product (94–97% pure) in the diet for 103 weeks(NTP 1986). Calculated dietary doses based on body weight and food intake were 0, 1,120, or2,240 mg/kg/day for male rats; 0, 1,200, or 2,550 mg/kg/day for female rats; 0, 3,200, or 6,650 mg/kg/dayfor male mice; and 0, 3,760, or 7,780 mg/kg/day for female mice. Animals were examined daily forclinical signs. Body weights and food consumption were measured throughout the study, and comprehensive gross and histological examinations were performed on all animals in all dose groups, including those that were moribund or died during the study. No hematology, clinical chemistry, or urineindices or thyroid hormone levels were evaluated. Liver degeneration and thrombosis were significantly(p<0.05) increased in male rats at 2,240 mg/kg/day; respective incidences in the control, low, and high dose groups were 13/50, 19/50, and 22/50 for degeneration and 1/50, 0/50, and 9/50 for thrombosis. Thethrombosis was characterized by a near total occlusion of a major hepatic blood vessel by a dense fibrin coagulum. Neoplastic nodules in the liver were significantly increased in a dose-related manner in males exposed to doses ≥1,120 mg/kg/day and in females exposed to 2,550 mg/kg/day. However, no treatment-related increases were observed in the incidence of hepatocellular carcinomas. Other effects in exposedrats included fibrosis of the spleen, lymphoid hyperplasia of the mandibular lymph nodes, and acanthosisof the forestomach at 2,240 mg/kg/day. In mice, histopathological changes occurred in males exposed to3,200 mg/kg/day, including centrilobular hypertrophy and granulomas in the liver and follicular cell hyperplasia in the thyroid. An MRL was not derived because the lowest tested dose, 1,120 mg/kg/day inmale rats, is a LOAEL for a liver lesion (neoplastic nodules) that is precancerous and associated withthrombosis in the same tissue. PBDEs 433. HEALTH EFFECTS3.1 INTRODUCTION The primary purpose of this chapter is to provide public health officials, physicians, toxicologists, and other interested individuals and groups with an overall perspective on the toxicology of PBDEs. Itcontains descriptions and evaluations of toxicological studies and epidemiological investigations andprovides conclusions, where possible, on the relevance of toxicity and toxicokinetic data to public health.A glossary and list of acronyms, abbreviations, and symbols can be found at the end of this profile.PBDEs are classes of brominated hydrocarbons that were previously used as flame retardant additives inplastics, textiles, and other materials. Production of PBDEs began in the 1970s and has continued untilrecently. PentaBDE and octaBDE mixtures were voluntarily withdrawn from the U.S. marketplace bytheir manufacturers at the end of 2004; however, the manufacture and use of decaBDE continued past thatdate (EPA 2010). In December of 2009, the two remaining U.S. producers of decaBDE and the largest U.S. importer of this product announced commitments to phase out manufacture and importation ofdecaBDE for most uses in the United States by December 31, 2012, and to end manufacture and importfor all uses by the end of 2013 (EPA 2013j). Although PBDEs are no longer produced or used, concern continues to exist for health effects of PBDEs due to evidence that PBDE congeners have become ubiquitously distributed in the environment and are present in tissues and breast milk of the generalpopulation (EPA 2010; Meijer et al. 2008; Park et al. 2011; Rawn et al. 2014; Schecter et al. 2010). PBDEs comprise compounds in which 1–10 bromine atoms are attached to the biphenyl structure in up to 209 different combinations. Based on the number of bromine substituents, there are 10 homologousgroups of PBDEs (monobrominated through decabrominated), each containing one or more isomers. PBDEs are structurally similar when viewed in one dimension, differing only in the ether linkage between the two phenyl rings in PBDEs, but the oxygen bridge confers three-dimensional conformationaldifferences that can influence toxicological properties. Consequently, on the basis of chemical structure, it cannot be assumed that the health effects of PBDE congeners are necessarily similar. Reviews on thehealth effects and other aspects of PBDEs include those by Bellinger (2013), Darnerud et al. (2001), deBoer et al. (2000a), de Wit (2002), Dingemans et al. (2011), EPA (2008a, 2008b, 2008c, 2008d), Gill etal. (2004), Hardy (1999, 2002a, 2002b), Hardy et al. (2009), Markowski (2007), McDonald (2002), Rahman et al. (2001), Silberhorn et al. (1990), and WHO (2006). Discussions of health effects aredivided into lower-brominated congeners and decaBDE due to important differences in pharmacokinetics and toxicity of decaBDE compared to lower-brominated PBDEs. Toxicity data for previously used PBDEs 443. HEALTH EFFECTSPBDE mixtures as well as individual PBDE congeners are included in this profile, with mixtures beingcategorized by their most prominent congener (see Table 4-3, Physical and Chemical Properties ofTechnical PBDE Mixtures). Using current health effects evaluation procedures, toxicity data forindividual congeners may over- or underestimate the actual health risk of PBDE mixtures because congeners vary in toxic potency and may be influenced by other congeners in an additive, less-thanadditive, or more-than-additive way. It is also important to recognize that the PBDEs to which peoplemay be exposed may be different from the original PBDE source because of possible changes in congenercomposition resulting from differential partitioning and transformation in the environment and/ordifferential biological metabolism and retention.3.2 DISCUSSION OF HEALTH EFFECTS BY ROUTE OF EXPOSURE To help public health professionals and others address the needs of persons living or working nearhazardous waste sites, the information in this section is organized first by route of exposure (inhalation, oral, and dermal) and then by health effect (death, systemic, immunological, neurological, reproductive, developmental, genotoxic, and carcinogenic effects). These data are discussed in terms of three exposure periods: acute (14 days or less), intermediate (15–364 days), and chronic (365 days or more).Levels of significant exposure for each route and duration are presented in tables and illustrated in figures. The points in the figures showing no-observed-adverse-effect levels (NOAELs) or lowest-observed-adverse-effect levels (LOAELs) reflect the actual doses (levels of exposure) used in the studies.LOAELs have been classified into "less serious" or "serious" effects. "Serious" effects are those thatevoke failure in a biological system and can lead to morbidity or mortality (e.g., acute respiratory distress or death). "Less serious" effects are those that are not expected to cause significant dysfunction or death,or those whose significance to the organism is not entirely clear. ATSDR acknowledges that a considerable amount of judgment may be required in establishing whether an end point should beclassified as a NOAEL, "less serious" LOAEL, or "serious" LOAEL, and that in some cases, there will be insufficient data to decide whether the effect is indicative of significant dysfunction. However, the Agency has established guidelines and policies that are used to classify these end points. ATSDRbelieves that there is sufficient merit in this approach to warrant an attempt at distinguishing between"less serious" and "serious" effects. The distinction between "less serious" effects and "serious" effects is considered to be important because it helps the users of the profiles to identify levels of exposure at whichmajor health effects start to appear. LOAELs or NOAELs should also help in determining whether or not PBDEs 453. HEALTH EFFECTSthe effects vary with dose and/or duration, and place into perspective the possible significance of these effects to human health. The significance of the exposure levels shown in the Levels of Significant Exposure (LSE) tables and figures may differ depending on the user's perspective. Public health officials and others concerned with appropriate actions to take at hazardous waste sites may want information on levels of exposure associated with more subtle effects in humans or animals (LOAELs) or exposure levels below which no adverse effects (NOAELs) have been observed. Estimates of levels posing minimal risk to humans (Minimal Risk Levels or MRLs) may be of interest to health professionals and citizens alike.A User's Guide has been provided at the end of this profile (see Appendix B). This guide should aid in the interpretation of the tables and figures for Levels of Significant Exposure and the MRLs.3.2.1 Inhalation Exposure A few studies have examined groups of chemical workers involved in the manufacture and distribution ofPBDEs (Bahn et al. 1980; Brown et al. 1981; Chanda et al. 1982; Landrigan et al. 1979; Rosenman et al. 1979; Stross et al. 1981). These people are believed to have been exposed predominantly by dermalcontact and inhalation, although the oral route cannot be ruled out. Results from these studies, therefore, are discussed in this section as well as in Section 3.2.3. The highest NOAEL and all LOAEL values fromeach reliable inhalation study of health effects end points in each species and duration category forPBDEs are recorded in Table 3-1 and plotted in Figure 3-1.3.2.1.1 Death No studies were located regarding death in humans after inhalation exposure to PBDEs.No deaths occurred in groups of five male and five female rats that were chamber-exposed to pentaBDEaerosol (compound dissolved in corn oil), octaBDE dust, or decaBDE dust in concentrations as high as200,000, 60,000, or 48,200 mg/m3, respectively, for 1 hour and observed for the following 14 days (IRDC 1974, 1975a, 1975b). Confidence in these studies is limited by a lack of control data. There was no mortality in rats that were exposed to dusts of commercial octaBDE products at concentrations of174 mg/m3 for 8 hours/day for 14 consecutive days (Great Lakes Chemical Corporation 1978) or≤202 mg/m3 for 6 hours/day, 5 days/week for 13 weeks (Great Lakes Chemical Corporation 2000).20633.7241651651650.63.71651651651653. HEALTH EFFECTS Table 3-1 Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - Inhalation PBDEs46Exposure/ Duration/a FrequencyKey to Species (Route)Figure (Strain) ACUTE EXPOSURE Systemic 1 Rat 14 d 8 hr/d(CD) System Resp CardioGastroHematoHepaticRenalEndocrOcularBd WtLOAEL NOAEL (mg/m³) Less Serious (mg/m³) Serious (mg/m³) 3.7 24 (reversible rapid breathing) 165 165 165 0.6 3.7 (hepatocytomegaly and focal hepatocellular degeneration) 165 165 165 165 Reference Chemical Form Comments Great Lakes Chemical Corporation 1978 OctaBDE (technical) 2358162022022022022021.1162021.1162022022022356162023. HEALTH EFFECTS Table 3-1 Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - Inhalation (continued) PBDEs47a Key to Figure Species (Strain) Exposure/ Duration/ Frequency (Route) INTERMEDIATE EXPOSURE Systemic 2 Rat (CD) 13 wk 5 d/wk 6 hr/d System Resp NOAEL (mg/m³) 16 Less Serious (mg/m³) 202 (alveolar histiocytosis, chronic active lung inflammation) LOAEL Serious (mg/m³) Reference Chemical Form Great Lakes Chemical Corporation 2000 OctaBDE (technical) Comments Cardio 202 Gastro 202 Hemato 202 Musc/skel 202 Hepatic 1.1 16 (centrilobular hepatocellular hypertrophy) Renal Endocr 202 b1.1 16 (decreased serum T4, increased serum TSH) Dermal 202 Ocular 202 Immuno/ Lymphoret Bd Wt 202 3 Rat (CD) 13 wk 5 d/wk 6 hr/d 16 202 (grossly discolored and enlarged bronchial and mediastinal lymph nodes associated with chronic Great Lakes Chemical Corporation 2000 OctaBDE (technical) active lung inflammation and alveolar histiocytosis) 2357162023. HEALTH EFFECTSTable 3-1 Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - Inhalation (continued) a Key to Figure Species (Strain) Exposure/ Duration/Frequency (Route)System NOAEL (mg/m³) LOAEL Reference Chemical Form Comments Less Serious (mg/m³) Serious (mg/m³) PBDEs48Reproductive 4 Rat 13 wk 16 F 202 F (absence of corpora lutea Great Lakes Chemical5 d/wk(CD) in ovaries) Corporation 20006 hr/d OctaBDE (technical) a The number corresponds to entries in Figure 3-1. b Used to derive an intermediate-duration (15-364 days) inhalation minimal risk level (MRL) of 0.006 mg/m3 for lower brominated diphenyl ethers. The MRL was derived by converting the animal NOAEL of 1.1 mg/m3 to a duration-adjusted human equivalent concentration (NOAELHEC) of 0.53 mg/m3, and dividing by an uncertainty factor of 30 (3 for species to species extrapolation with dosimetric adjustments and 10 for human variability) and a modifying factor of 3 (for an incomplete data base). Note on chemical form: The chemical form in all studies was a technical octaBDE mixture (exact composition was not reported). Bd Wt = body weight; Cardio = cardiovascular; d = day(s); Endocr = endocrine; F = female; Gastro = gastrointestinal; Hemato = hematological; hr = hour(s); Immuno/Lymphoret = immunological/lymphoreticular; LOAEL = lowest observed adverse effect level; Musc/skel = muscular/skeletal; NOAEL = no observed adverse effect level; Resp = respiratory; T4 = thyroxine; TSH = thyroid stimulating hormone; wk = week(s) Respiratory Cardiovascular Gastrointestinal Hematological Hepatic Renal Endocrine Ocular Body Weight mg/m3 1000 1r 1r 1r 1r 1r 1r 1r 100 1r 10 1r 1r 1 1r 0.1 c-Cat d-Dogr-Rat p-Pigq-Cow k-Monkeym-Mouse h-Rabbit a-Sheep f-Ferret j-Pigeone-Gerbil s-Hamster g-Guinea Pig n-Mink o-Other Cancer Effect Level-Animals LOAEL, More Serious-AnimalsLOAEL, Less Serious-AnimalsNOAEL - Animals Cancer Effect Level-Humans LOAEL, More Serious-HumansLOAEL, Less Serious-HumansNOAEL - Humans LD50/LC50Minimal Risk Level for effects other than Cancer 3. HEALTH EFFECTS Figure 3-1. Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - Inhalation Acute (≤14 days) Systemic PBDEs49RespiratoCardiovaGastrointHematoloMusculosHepatic Renal EndocrinDermal Ocular Body WeiImmuno/Reprodu1000 mg/m3 2r 2r 2r 2r 2r 2r 2r 2r 2r 3r 4r 100 2r 2r 2r 3r 4r 10 2r 2r1 0.1 0.01 0.001 c-Cat d-Dogr-Rat p-Pigq-Cow k-Monkeym-Mouse h-Rabbit a-Sheep f-Ferret j-Pigeone-Gerbil s-Hamster g-Guinea Pig n-Mink o-Other Cancer Effect Level-Animals LOAEL, More Serious-AnimalsLOAEL, Less Serious-AnimalsNOAEL - Animals Cancer Effect Level-Humans LOAEL, More Serious-HumansLOAEL, Less Serious-HumansNOAEL - Humans LD50/LC50Minimal Risk Level for effects other than Cancer ry scular estinal gical keletal e ght Lymphor ctive 3. HEALTH EFFECTS Figure 3-1. Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - Inhalation (Continued)Intermediate (15-364 days)SystemicPBDEs50 PBDEs 513. HEALTH EFFECTS3.2.1.2 Systemic EffectsSystemic effects that have been observed in humans and animals following inhalation exposure to PBDEsare described below.Respiratory Effects. No studies were located regarding respiratory effects in humans after inhalationexposure to PBDEs.Transient signs of respiratory distress that included tachypnea or dyspnea developed in rats that were chamber-exposed to pentaBDE aerosol (compound dissolved in corn oil), octaBDE dust, or decaBDEdust in very high concentrations of 200,000, 60,000, and 48,200 mg/m3, respectively, for 1 hour (IRDC 1974, 1975a, 1975b). Confidence in these effect levels is low due to a small number of tested animals and lack of control data. One 14-day inhalation study of commercial octaBDE has been conducted. In this study, rats werechamber-exposed to concentrations of 0, 0.6, 3.7, 23.9, or 165.2 mg/m3 as powdered dust for 8 hours/dayfor 14 consecutive days (Great Lakes Chemical Corporation 1978). Increased respiration rate occurred at≥23.9 mg/m3. The rapid breathing pattern developed by the end of each exposure period, alwaysdisappeared by the following morning, and was not observed at lower exposure concentrations. Histological examinations of the control and 165.2 mg/m3 rats (other groups not examined) showed no changes in tissues that included nasal turbinates, trachea, lungs, and mediastinal lymph nodes). Histological changes in the lungs, but no clearly observed changes in the nasal cavity, were found in astudy of rats that were nose-only exposed to 0, 1.1, 16, or 202 mg/m3 as dust aerosol for 6 hours/day, 5 days/week for 13 weeks (Great Lakes Chemical Corporation 2000). The pulmonary effects included alveolar histiocytosis and chronic active inflammation, which occurred in both sexes, and were onlyclearly induced at 202 mg/m3. Total incidences of alveolar histiocytosis in the 0, 1.1, 16, and 202 mg/m3 exposure groups were 3/10, 5/10, 5/10, and 10/10 in males, respectively, and 0/10, 5/10, 2/10, and 10/10 in females, respectively. Respective total incidences of chronic active lung inflammation were 0/10, 0/10, 2/10, and 10/10 in males, and 0/10, 1/10, 1/10, and 10/10 in females. Both lesions werepredominantly minimal or mild in severity, with moderate severity occurring in a few high-dose animals.Additional effects included gross pulmonary changes in both sexes at 202 mg/m3; these included lungfirmness and white discoloration and/or enlargement in the bronchial and/or mediastinal lymph nodes. The gross lymph node changes correlated with the histological granulomatous inflammation. Effects in PBDEs 523. HEALTH EFFECTSnasal tissues were equivocal. Incidences of nasal goblet cell hypertrophy were slightly increased in nasallevel II of both sexes at ≥1.1 mg/m3, but changes in incidence were not clearly dose-related and there was essentially no increase in severity from minimal levels with increasing dose. Total incidences of gobletcell hypertrophy in nasal level II in the 0, 1.1, 16, and 202 mg/m3 exposure groups were 4/10, 9/10, 6/10, and 10/10 respectively, in males, and 2/10, 6/10, 4/10, and 8/10, respectively, in females. Minimalseverity goblet cell hypertrophy was also slightly increased in nasal level IV in males at 202 mg/m3 (4/10, 0/10, 1/10, and 8/10), but not in females.Cardiovascular Effects. No studies were located regarding cardiovascular effects in humans afterinhalation exposure to PBDEs.No histopathological changes were observed in the heart of rats that were exposed to dusts of commercialoctaBDE products at concentrations of 174 mg/m3 for 8 hours/day for 14 consecutive days (Great Lakes Chemical Corporation 1978), or ≤202 mg/m3 for 6 hours/day, 5 days/week for 13 weeks (Great Lakes Chemical Corporation 2000).Gastrointestinal Effects. No studies were located regarding gastrointestinal effects in humans afterinhalation exposure to PBDEs.No histopathological changes were observed in the stomach and lower gastrointestinal tract of rats thatwere exposed to dusts of commercial octaBDE products at concentrations of 174 mg/m3 for 8 hours/dayfor 14 consecutive days (Great Lakes Chemical Corporation 1978), or ≤202 mg/m3 for 6 hours/day, 5 days/week for 13 weeks (Great Lakes Chemical Corporation 2000). Hematological Effects. No studies were located regarding hematological effects in humans afterinhalation exposure to PBDEs.No adverse hematological changes occurred in rats that were exposed to 24.4 or 174 mg/m3 ofcommercial octaBDE dust aerosol for 8 hours/day for 14 consecutive days (Great Lakes ChemicalCorporation 1978). Evaluation of a limited number of indices (hemoglobin, hematocrit, total erythrocytecount, and total and differential leukocyte counts) showed no unusual responses except for an elevation inleukocyte numbers. The observed increase in leukocyte counts was considered to be an unusual responseby the investigators, although it was within the normal range for control rats in their laboratory. Comprehensive hematological assessments showed no unusual changes in rats exposed to commercial PBDEs 533. HEALTH EFFECTSoctaBDE as dust aerosol at concentrations of ≤202 mg/m3 for 6 hours/day, 5 days/week for 13 weeks(Great Lakes Chemical Corporation 2000).Hepatic Effects. No studies were located regarding hepatic effects in humans after inhalationexposure to PBDEs.Hepatic effects were observed in a 14-day inhalation study of dusts of commercial octaBDE mixtures. Inthis study, rats were chamber-exposed to concentrations of 0, 0.6, 3.7, 23.9, or 165.2 mg/m3 as powdered dust for 8 hours/day for 14 consecutive days (Great Lakes Chemical Corporation 1978). Increased liverweight and hepatic histological changes occurred in rats exposed to concentrations ≥3.7 mg/m3. At3.7 mg/m3, the liver lesions consisted of very slight to slight severity focal to multifocal cytoplasmic enlargement of the hepatocytes, accompanied by focal acidophilic degeneration of individual to smallgroups of cells. The liver lesions were similar in rats exposed to concentrations ≥24.4 mg/m3, except thatthe hepatocyte enlargement was multifocal to diffuse in distribution and accompanied by focal, small to large areas of hepatocellular necrosis of very slight to marked degree.Similar hepatic changes were found in a study of rats that were nose-only exposed to 0, 1.1, 16, or202 mg/m3 commercial octaBDE as dust aerosol for 6 hours/day, 5 days/week for 13 weeks (Great LakesChemical Corporation 2000). The liver was affected in both sexes as shown by dose-related increases incentrilobular hepatocellular hypertrophy at ≥16 mg/m3 and increased liver weight (absolute and relative)at 202 mg/m3. Respective total incidences of centrilobular hepatocellular hypertrophy (predominantlyminimal to mild) in the 0, 1.1, 16, and 202 mg/m3 groups were 1/10, 0/10, 3/10, and 10/10 in males, and 0/10, 0/10, 3/10, and 6/10 in females. Serum chemistry evaluations showed no clear effects of exposure. Serum cholesterol was significantly increased (66.2% more than controls, p<0.01) in 202 mg/m3 females,but the magnitude of the elevation was not considered toxicologically significant. Some other statisticallysignificant serum chemistry alterations (increased mean globulin and total protein, decreasedalbumin/globulin ratio) also occurred in females exposed to 202 mg/m3, but these changes were notconsidered exposure-related due to small magnitudes of changes and lack of similar changes in the males.Renal Effects. No studies were located regarding renal effects in humans after inhalation exposure to PBDEs.No histopathological changes were observed in the kidneys or urinary bladder of rats that were exposed todusts of commercial octaBDE products at concentrations of 174 mg/m3 for 8 hours/day for 14 consecutive PBDEs 543. HEALTH EFFECTSdays (Great Lakes Chemical Corporation 1978) or ≤202 mg/m3 for 6 hours/day, 5 days/week for13 weeks (Great Lakes Chemical Corporation 2000). Urinalyses were not performed in any of thesestudies.Endocrine Effects. There is evidence suggestive of hypothyroidism in a small group of workers who were occupationally exposed to decaBDE (Bahn et al. 1980). In another study, plasma levels of thyroid hormones (T3 and free T4) and eight PBDE congeners (tetra- to heptaBDEs) were monitored for 198– 221 days in three electronic dismantling workers (Pettersson et al. 2002). The hormones remained withinnormal ranges and there were no correlations between levels of hormones and the plasma concentrations of congeners. An acute inhalation study of commercial octaBDE dust in rats showed no histopathological changes in thethyroids, parathyroids, adrenals, or pituitary following chamber exposure to 174 mg/m3 as powdered dustfor 8 hours/day for 14 consecutive days (Great Lakes Chemical Corporation 1978). Rats that were nose-only exposed to commercial octaBDE at concentrations of 1.1, 16, or 202 mg/m3 for 6 hours/day, 5 days/week for 13 weeks similarly showed no histological changes in the adrenals, pancreas,parathyroids, pituitary, or thyroids (Great Lakes Chemical Corporation 2000). Measurements of serumlevels of thyroid hormones in the 13-week rat study, however, showed exposure-related decreases inmean thyroxine (total T4) in both sexes exposed at ≥16 mg/m3, and increases in TSH in males exposed at ≥16 mg/m3 and in females exposed at 202 mg/m3. The changes were usually statistically significant(p<0.05 or p<0.01) compared to controls and were considered by the investigators to be consistent with chemical-induced hypothyroidism. There were no serum T3 changes, thyroid-attributable clinical signs orbody weight effects, or gross or histopathological changes in the thyroid. The 1.1 mg/m3 LOAEL forthyroid effects was used as the basis for the intermediate-duration MRL for inhalation exposure to octaBDE, as indicated in the footnote to Table 3-1 and discussed in Chapter 2 and Appendix A.Dermal Effects. No studies were located regarding dermal effects in humans after inhalation exposure to PBDEs.No gross or histological changes in the skin were observed in rats that were nose-only exposed tocommercial octaBDE as dust aerosol at concentrations of ≤202 mg/m3 for 6 hours/day, 5 days/week for13 weeks (Great Lakes Chemical Corporation 2000). PBDEs 553. HEALTH EFFECTSOcular Effects. No studies were located regarding ocular effects in humans after inhalation exposure to PBDEs.Transient signs of ocular irritation that included eye squint, erythema, and/or ocular discharge were observed in rats that were chamber-exposed to pentaBDE aerosol (compound dissolved in corn oil), octaBDE dust, or decaBDE dust in concentrations of 2,000, 2,000, and 48,200 mg/m3, respectively, for1 hour (IRDC 1974, 1975a, 1975b). Confidence in these effect levels is low due to a small number oftested animals and lack of control data.No histopathological changes were observed in eyes of rats that were chamber-exposed to ≤174 mg/m3 ofcommercial octaBDE as powdered dust for 8 hours/day for 14 consecutive days (Great Lakes ChemicalCorporation 1978). Opthalmoscopic and histological examinations showed no ocular effects in ratsfollowing nose-only exposure to ≤202 mg/m3 of commercial octaBDE dust aerosol for 6 hours/day, 5 days/week for 13 weeks (Great Lakes Chemical Corporation 2000).3.2.1.3 Immunological and Lymphoreticular Effects A study conducted in China examined the association between serum levels of four BDEs (28, 47, 66, and 209), as well as PCBs and organochlorine pesticides, and the prevalence of asthma in children (3–6 years old) living in a heavily populated area (Meng et al. 2016). The study involved 620 cases and 218 controls. Serum analyses showed significantly higher levels of BDEs in cases than in controls.BDE 209 had the highest mean concentration in cases, 4.02 ng/g lipid, compared to 1.73 ng/g lipid in controls. In multivariable-adjusted analyses, the odds ratio (OR) for risk of asthma associated with 1 ng/gincrease was statistically significant only for BDE 28 (OR 3.63; 95% confidence interval [CI] 1.23– 10.70). Strafication of children based on asthma severity showed that BDE 209 was positively correlated with the severity of the condition (OR 1.40; 95% CI 1.14–1.72). PCBs and organochlorine pesticide alsowere associated with increased risk of asthma.No histopathological changes were observed in the spleen, mesenteric or mediastinal lymph nodes, orbone marrow from rats that were exposed to 174 mg/m3 of octaBDE dust for 8 hours/day for14 consecutive days (Great Lakes Chemical Corporation 1978). Rats that were nose-only exposed to commercial octaBDE at concentrations of 1.1, 16, or 202 mg/m3 as dust aerosol for 6 hours/day, 5 days/week for 13 weeks similarly showed no effects in bone marrow, spleen, or thymus, although grosschanges in pulmonary lymph nodes were observed at 202 mg/m3 (Great Lakes Chemical Corporation PBDEs 563. HEALTH EFFECTS2000). The effects included discolored and/or enlarged bronchial and mediastinal lymph nodes, and appeared to be associated with concurrent granulomatous inflammation of the lungs. 3.2.1.4 Neurological Effects No studies were located regarding neurological effects in humans after inhalation exposure to PBDEs.No clinical signs of neurotoxicity were observed in rats that were exposed to dusts of commercialoctaBDE products at concentrations of 174 mg/m3 for 8 hours/day for 14 consecutive days (Great Lakes Chemical Corporation 1978) or ≤202 mg/m3 for 6 hours/day, 5 days/week for 13 weeks (Great Lakes Chemical Corporation 2000). Histological examinations of nervous system tissues, performed only in the 13-week study, showed no effects in the brain (forebrain, midbrain, hindbrain), optic nerve, or a peripheral nerve (sciatic).3.2.1.5 Reproductive EffectsNo studies were located regarding reproductive effects in humans after inhalation exposure to PBDEs.No histopathological changes were observed in testes or ovaries from rats that were exposed to commercial octaBDE at concentrations ≤174 mg/m3 as powdered dust for 8 hours/day for 14 consecutive days (Great Lakes Chemical Corporation 1978). A histological effect in the ovaries was found in a studyof rats that were nose-only exposed to 0, 1.1, 16, or 202 mg/m3 as dust aerosol for 6 hours/day, 5 days/week for 13 weeks (Great Lakes Chemical Corporation 2000). Absence of corpora lutea, based on qualitative evaluation of step sections of the ovary, was found in 3/10 females at 202 mg/m3, compared to 0/10 incidences in the control and both lower exposure groups. The investigators interpreted this 30%increase in incidence be treatment-related because an absence of corpora lutea was considered unusual in rats at 20 weeks of age. No gross or histopathological changes were observed in the oviduct, uterus, orvagina, or in male reproductive tissues (testes with epididymides and vas deferens).3.2.1.6 Developmental Effects No studies were located regarding developmental effects in humans or animals after inhalation exposureto PBDEs. PBDEs 573. HEALTH EFFECTS3.2.1.7 CancerNo studies were located regarding cancer in humans or animals after inhalation exposure to PBDEs.3.2.2 Oral ExposureHuman data presented below are primarily from studies that attempted to identify relationships betweenconcentrations of PBDE in serum or other tissues and various health outcomes. Details of PBDEexposure in these study populations are typically unknown. However, exposure is presumed to have beenprimarily by the oral route for those studies presented below.The highest NOAEL and all LOAEL values from each reliable study of health effects end points in eachspecies and duration category for PBDEs are recorded in Tables 3-2 (lower PBDEs) or 3-3 (decaBDE)and plotted in Figures 3-2 (lower PBDEs) or 3-3 (decaBDE).3.2.2.1 Death Single-dose gavage LD50 values of 5,000 and 6,200 mg/kg were determined for pentaBDE (Saytex 115 and DE-71, respectively) in rats that were observed for 14 days (British Industrial Biological ResearchAssociation 1977; Pharmakon Research International Inc. 1984). Another study found that a single5,000 mg/kg dose of pentaBDE caused deaths in four of five rats in the 14 days following treatment,whereas doses ≤500 mg/kg caused no mortality (IRDC 1975b). No deaths occurred in rats exposed to pentaBDE in estimated dietary doses of ≤90 mg/kg/day for 28 days (IRDC 1976) or ≤100 mg/kg/day for90 days (WIL Research Laboratories 1984).No deaths occurred in rats that were administered octaBDE by gavage in single doses ≤5,000 mg/kg and observed for the following 14 days (IRDC 1975a). Intermediate-duration dietary studies with octaBDE, resulted in no mortality in rats exposed to estimated dietary doses of ≤90 mg/kg/day for 28 days or≤750 mg/kg/day for 13 weeks (IRDC 1976, 1977).No deaths occurred in rats that were treated with a single gavage dose of ≤5,000 mg/kg of decaBDE or≤2,000 mg/kg of 77.4% decaBDE (containing 21.8% nonaBDE and 0.8% octaBDE) and observed for thefollowing 14 days (IRDC 1974; Norris et al. 1975a). No mortality was observed in rats and mice thatwere exposed to decaBDE via diet in estimated doses of ≤16,000 and ≤19,000 mg/kg/day, respectively,for 14 days (NTP 1986). 21066200209050002107500021191.20.621211.20.61.20.62071101003. HEALTH EFFECTSTable 3-2 Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - Oral a Key to Figure Species (Strain) Exposure/ Duration/ Frequency (Route) ACUTE EXPOSURE Death 1 Rat (Wistar) once (GO) System NOAEL (mg/kg) Less Serious (mg/kg) LOAEL Serious (mg/kg) 6200 (44-day LD50) Reference Chemical Form British Industrial Biological Research Association 1977 PentaBDE (DE-71) Comments 2 Rat Spartan once (GO) 5000 (4/5 died) IRDC 1975b PentaBDE (technical) 3 Rat (Sprague-Dawley) Systemic 4 Rat (Sprague-Dawley) once (GO) once (GO) Hepatic 1.2 M 5000 (14-day LD50) Pharmakon Research International Inc. 1984 PentaBDE (Saytex 115) Albina et al. 2010 PentaBDE (BDE99) Renal 0.6 M (phagolysosomes in renal tubules) 5 Rat (Sprague-Dawley) once (GO) Hepatic 1.2 M Alonso et al. 2010 PentaBDE (BDE99) No biologically relevant changes in hepatic serum chemistry. Renal 0.6 M 1.2 M (increased total protein in urine) Endocr 0.6 M (reduced serum testosterone) 6 Rat (CD) 10 d Gd 6-15 (GO) Bd Wt 10 100 (20% reduction in maternal body weight gain) Argus Research Laboratories 1985a PentaBDE (Saytex 115) PBDEs582132221448 4020021458200215240 20040200215340 200402003. HEALTH EFFECTSTable 3-2 Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - Oral (continued) PBDEs59a Key to Species Figure (Strain) 7 Rat (Sprague-Dawley) 14 d Gd 6-19 (GO) 8 Rat (Wistar) 7 d 1 x/d (GO) 9 Rat (Wistar) 14 d 1 x/d (GO) 10 Rat (Wistar) 7 d 1 x/d 11 Rat (Wistar) 14 d 1 x/d Exposure/ Duration/ Frequency (Route) System Bd Wt Hepatic Bd Wt Hepatic Bd Wt Hepatic Bd Wt Hepatic Bd Wt LOAEL NOAEL (mg/kg/day) Less Serious (mg/kg/day) Serious (mg/kg/day) Reference Chemical Form Comments 2 F Blanco et al. 2012 PentaBDE (BDE99) 8 F 200 F 200 F 40 F 40 F 40 F (porphyria) 8 F (porphyria) 200 F (fatty degeneration) 200 F (7% decrease in body weight gain) Bruchajzer 2011 PentaBDE (technical) Bruchajzer 2011 PentaBDE (technical) Bruchajzer et al. 2010 PentaBDE (technical) 40 F 40 F 200 F (fatty degeneration) 200 F (10% decrease in body weight gain) Bruchajzer et al. 2010 PentaBDE (technical) 2156Table 3-2 Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - Oral (continued) Exposure/ LOAEL Reference Chemical Form Comments Duration/ Frequency NOAEL Less Serious Serious (Route) System (mg/kg) (mg/kg) (mg/kg) once Hepatic 200 F 2000 F (fatty degeneration) 5 d observation 200 2000Bruchajzer et al. 2011 (GO) PentaBDE (technical) Bd Wt 200 F 2000 F (11% decrease in body 200 weight gain) 2000200 F 2000 F (11% decrease in body 200 weight gain) 20007 d Hepatic 2 F 8 F (porphyria) 1 x/d 2 8(GO) Bruchajzer et al. 2012 OctaBDE (technical) Bd Wt 8 F 40 F (9% decrease in body 8 weight gain) 4014 d Hepatic 8 F 40 F (porphyria) 1 x/d 8 40(GO) Bruchajzer et al. 2012 OctaBDE (technical) Bd Wt 8 F 40 F (5% decrease in body 8 weight gain) 4014 d Hepatic 76.6 M 1 x/d 76.6(GO) Carlson, 1980b OctaBDE (technical) 14 d Hepatic 56.4 M 1 x/d 56.4(GO) Carlson, 1980b PentaBDE (technical) 21572158204220433. HEALTH EFFECTSPBDEs60a Key to Species Figure (Strain) 12 Rat (Wistar) 13 Rat (Wistar) 14 Rat (Wistar) 15 Rat (Sprague-Dawley) 16 Rat (Sprague-Dawley) 2109182180120120219861818206818183622051414143. HEALTH EFFECTSTable 3-2 Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - Oral (continued) PBDEs61a Key to Figure 17 18 19 20 21 Species (Strain) Rat (Sprague-Dawley) Rat (Sprague-Dawley) Rat (Sprague-Dawley) Rat (Sprague-Dawley) Rat (Sprague-Dawley) Exposure/ Duration/ Frequency (Route) 14 d 1 x/d (GO) 6 d Gd 6.5-11.5 (GO) 14 d 1 x/d (G) 14 d 1 x/d (GO) 2 wk 1 x/d (GO) System Endocr Endocr Bd Wt Endocr Bd Wt Hepatic Endocr Bd Wt Endocr Bd Wt Metab NOAEL (mg/kg/day) 120 F 120 F 6 F 18 F 36 F 14 M 14 M LOAEL Less Serious Serious (mg/kg/day) (mg/kg/day) 18 (reduced serum T4) 18 F (reduced serum T4, reduced T4 protein binding) 18 F (reduced liver vitamin A) 18 F (reduced serum T4) 14 M (reduced serum T4) Reference Chemical Form Darnerud and Sinjari 1996 PentaBDE (Bromkal 70) Ellis-Hutchings et al. 2009 PentaBDE (DE-71) Hallgren and Darnerud 2002 TetraBDE (BDE47) Hallgren et al. 2001 PentaBDE (Bromkal 70-5DE) Hoppe and Carey 2007 PentaBDE (technical) Comments No exposure-related changes in serum thyroid hormone levels. No exposure-related changes in fat pad weight, adipocyte number, size, viability, lipolysis or glucose oxidation. 20895000209150022150.0620752522903 306022933 3060229730 60603. HEALTH EFFECTSTable 3-2 Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - Oral (continued) a Key to Figure Species (Strain) Exposure/ Duration/ Frequency (Route) System NOAEL (mg/kg) Less Serious (mg/kg) LOAEL Serious (mg/kg) Reference Chemical Form Comments 22 23 24 Rat Spartan Rat Spartan Rat (Wistar) once (GO) once (GO) once Gd 6 (GO) Bd Wt Bd Wt Endocr 5000 500 b 0.06 F (reduced serum T4) IRDC 1975a OctaBDE (technical) IRDC 1975b PentaBDE (technical) Kuriyama et al. 2007 PentaBDE (BDE99) 25 26 Rat (CD) Rat (Wistar) 10 d Gd 6-15 (GO) 5 d Pnd 22-26 (GO) Bd Wt Endocr Bd Wt 25 3 F 60 F 30 F (reduced serum T4) Life Science Research Israel Ltd. (1987) OctaBDE (FR-1208) Stoker et al. 2004 PentaBDE (DE-71) 27 Rat (Wistar) 5 d Pnd 23-27 (GO) Endocr Bd Wt 3 M 60 M 30 M (reduced serum T4) Stoker et al. 2004 PentaBDE (DE-71) 28 Rat (Wistar) 3 d 1 x/d (GO) Endocr Bd Wt 30 M 60 M 60 M (increased serum LH) Stoker et al. 2005 PentaBDE (DE-71) PBDEs62229924023040.720732550208110 3030020823 103002110183. HEALTH EFFECTSTable 3-2 Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - Oral (continued) PBDEs63a Key to Species Figure (Strain) 29 Rat (Wistar) 30 Rat (Wistar) 31 Rat (CD) 32 Rat (Long-Evans) 33 Rat (Long-Evans) 34 Mouse (C57BL/6N) Exposure/ Duration/ Frequency (Route) System NOAEL (mg/kg/day) Less Serious (mg/kg/day) LOAEL Serious (mg/kg/day) Reference Chemical Form Comments 9 d 1 x/d (GO) Bd Wt 240 M Stoker et al. 2005 PentaBDE (DE-71) Hershberger Assay (castrated rats, supplemented with s.c. testosterone) once Gd 6 (GO) Bd Wt 0.7 F Talsness et al. 2008 TetraBDE (BDE47) 10 d Gd 6-15 (GO) Bd Wt 25 50 (40% reduced maternal body weight gain) WIL Research Laboratories 1986 OctaBDE (DE-79) 4 d 1 x/d (GO) Endocr Bd Wt 10 F 300 F 30 F (reduced serum T4) Zhou et al. 2001 PentaBDE (DE-71) 4 d 1 x/d (GO) Endocr Bd Wt 3 F 300 F 10 F (reduced serum T4) Zhou et al. 2001 OctaBDE (DE-79) 14 d 1 x/d (GO) Endocr 18 (reduced serum T4) Darnerud and Sinjari 1996 PentaBDE (Bromkal 70) 2059500100 5005002061721872206918 361836211318183. HEALTH EFFECTSTable 3-2 Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - Oral (continued) a Key to Figure Species (Strain) Exposure/ Duration/ Frequency (Route) System NOAEL (mg/kg) Less Serious (mg/kg) LOAEL Serious (mg/kg) Reference Chemical Form Comments 35 Mouse C57BL/6J once (GO) Hepatic Endocr Bd Wt 500 F 100 F 500 F 500 F (reduced serum T4) Fowles et al. 1994 PentaBDE (DE-71) 36 Mouse (C57BL/6N) 14 d 1 x/d (GO) Hepatic Endocr Bd Wt 72 F 72 F 18 F (reduced serum T4) Fowles et al. 1994 PentaBDE (DE-71) 37 Mouse (C57BL/6N) 14 d 1 x/d (GO) Hepatic Endocr Bd Wt 18 F 36 F 36 F 18 F (reduced liver vitamin A) (reduced serum T4) Hallgren et al. 2001 PentaBDE (Bromkal 70-5DE) 38 Mouse (C57BL/6N) 14 d 1 x/d (GO) Endocr Bd Wt 18 F 18 F (reduced serum T4) Hallgren et al. 2001 TetraBDE (BDE47) PBDEs64224110 100100206636206718362060500206236723. HEALTH EFFECTSTable 3-2 Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - Oral (continued) LOAEL System NOAEL (mg/kg/day) Less Serious (mg/kg/day) Serious (mg/kg/day) PBDEs65a Key to Species Figure (Strain) 39 Mouse (C57BL/6) Immuno/ Lymphoret 40 41 42 43 Rat (Sprague-Dawley) Mouse (C57BL/6N) Mouse (C57BL/6N) Mouse (C57BL/6N) Exposure/ Duration/ Frequency (Route) Endocr 10 F 100 F (reduced serum T4) Bd Wt 100 F 36 18 36 (reduced in vitro production of IgG in mitogen- stimulated splenocytes) 500 F 36 F 72 F (reduced antibody response to sheep red blood cells, decreased thymus weight) Reference Chemical Form Richardson et al. 2008 TetraBDE (BDE47) Darnerud and Thuvander 1998 PentaBDE (Bromkal 70-5DE) Darnerud and Thuvander 1998 PentaBDE (DE-71) Fowles et al. 1994 PentaBDE (DE-71) Fowles et al. 1994 PentaBDE (DE-71) Comments No exposure-related changes in spleen or thymus weight, number or distribution of lymphocyte subpopulations, or in vitro IgG production. No exposure-related change in spleen or thymus weights or antibody respose to sheep red blood cells. 4 d 1 x/d (GO) 14 d 1 x/d (GO) 14 d 1 x/d (GO) once (GO) 14 d 1 x/d (GO) 21141821261.221200.621312229630 602298303. HEALTH EFFECTSTable 3-2 Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - Oral (continued) PBDEs66a Key to Species Figure (Strain) 44 Mouse (C57BL/6N) Neurological 45 Rat (Sprague-Dawley) Reproductive 46 Rat (Sprague-Dawley) 47 Rat (Sprague-Dawley) 48 Rat (Wistar) 49 Rat (Wistar) Exposure/ Duration/ Frequency (Route) System NOAEL (mg/kg/day) Less Serious (mg/kg/day) LOAEL Serious (mg/kg/day) Reference Chemical Form 14 d 1 x/d (GO) 18 F Hallgren et al. 2001 TetraBDE (BDE47) once (GO) 1.2 M Belles et al. 2010 PentaBDE (BDE99) once (GO) 0.6 F (decreased serum testosterone) Alonso et al. 2010 PentaBDE (BDE99) 14 d Gd 6-19 (GO) 2 F Blanco et al. 2012 PentaBDE (BDE99) 3 d 1 x/d (GO) 9 d 1 x/d (GO) 30 M 60 M (increased serum LH) 30 M (20% decrease in ventral prostate weight) Stoker et al. 2005 PentaBDE (DE-71) Stoker et al. 2005 PentaBDE (DE-71) Comments No exposure-related changes in spleen or thymus weight No changes in brain histology or in functional observation battery, open-field testing, passive avoidance test, or Morris water maze. No treatment-related changes in gravid uterine weight or number of implantation or resorptions. Hershberger Assay (castrated rats, supplemented with s.c. testosterone) 23030.722323002070200205810252130123. HEALTH EFFECTSTable 3-2 Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - Oral (continued) PBDEs67a Key to Species Figure (Strain) 50 Rat (Wistar) 51 Mouse (BALB/c) Developmental 52 Rat (CD) 53 Rat (CD) 54 Rat (Sprague-Dawley) Exposure/ Duration/ Frequency (Route) System NOAEL (mg/kg) Less Serious (mg/kg) LOAEL Serious (mg/kg) Reference Chemical Form once Gd 6 (GO) 0.7 F Talsness et al. 2008 TetraBDE (BDE47) 3 d 1 x/d (GO) 300 F Mercado-Feliciano and Bigsby 2008a PentaBDE (DE-71) 10 d Gd 6-15 (GO) 200 Argus Research Laboratories 1985a PentaBDE (Saytex 115) 10 d Gd 6-15 (GO) 10 25 (increased resorptions and reduced fetal body weight) Argus Research Laboratories 1985b OctaBDE (Saytex 115) 14 d Gd 6-19 (GO) 1 2 (delayed ossification, liver and heart hypertrophy) Blanco et al. 2012 PentaBDE (BDE99) Comments No dose-related changes in ovary weight or histology. Mice were ovariectomized 3 weeks prior to exposure; no change in uterus weight or uterus/vaginal histology. No changes in number, sex, and weight of fetuses, fetal death, early and late resorptions, gross malformations, or skeletal or visceral abnormalities. 21763021781202179120219912200122130.063. HEALTH EFFECTSTable 3-2 Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - Oral (continued) a Key to Figure Species (Strain) Exposure/ Duration/ Frequency (Route) System NOAEL (mg/kg/day) Less Serious (mg/kg/day) LOAEL Serious (mg/kg/day) Reference Chemical Form Comments 55 Rat (Long- Evans) 7 d Pnd 6-12 (GO) 30 M (impaired learning in visual discrimination task at Pnd 30-83) Dufault et al. 2005 PentaBDE (DE-71) 56 Rat (Sprague-Dawley) 6 d Gd 6.5-11.5 (GO) 120 Ellis-Hutchings et al. 2009 PentaBDE (DE-71) No exposure-related changes in embryo viability, growth, or morphology. 57 Rat (Sprague-Dawley) 14 d Gd 6.5-19.5 (GO) 120 Ellis-Hutchings et al. 2009 PentaBDE (DE-71) No exposure-related effects on fetal survival, growth, or malformations. 58 Rat (Sprague-Dawley) once Pnd 10 (GO) 1 (impaired learning and memory at 2 months) He et al. 2009 TetraBDE (BDE47) 59 Rat (Sprague-Dawley) once Pnd 10 (GO) 1 (impaired learning and memory and 23% decrease in relative uterine weight at 2 months) He et al. 2011 TetraBDE (BDE47) 60 Rat (Wistar) once Gd 6 (GO) b 0.06 (increased activity and impaired spermatogenesis in adult offspring) Kuriyama et al. 2005 PentaBDE (BDE99) PBDEs6822140.060.320742.51022460.0623020.1422510.883. HEALTH EFFECTSTable 3-2 Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - Oral (continued) PBDEs69a Key to Species Figure (Strain) 61 Rat (Wistar) 62 Rat (CD) 63 Rat (Wistar) 64 Rat (Wistar) 65 Rat (Sprague-Dawley) Exposure/ Duration/ Frequency (Route) System NOAEL (mg/kg) Less Serious (mg/kg) LOAEL Serious (mg/kg) Reference Chemical Form Comments once Gd 6 (GO) 0.06 0.3 (reduced serum T4 in Pnd 22 offspring) Kuriyama et al. 2007 PentaBDE (BDE99) 10 d Gd 6-15 (GO) 2.5 10 (minimal increased post-implantation loss) Life Science Research Israel Ltd. (1987) OctaBDE (FR-1208) once Gd 6 (GO) b 0.06 F (ultrastructural changes in ovaries in F1 females at Pnd 90, increased resorptions in F1 females mated to unexposed males) Talsness et al. 2005 PentaBDE (BDE99) once Gd 6 (GO) 0.14 F (reduced number of secondary ovarian follicles at Pnd 38 and ultrastructural changes in the ovary at Pnd 100 in offspring) Talsness et al. 2008 TetraBDE (BDE47) once Pnd 10 (G) 0.8 M 8 M (decreased spontaneous activity, impaired habituation) Viberg et al. 2005 PentaBDE (BDE99) 20722550211111023691021726.821860.821870.710.53. HEALTH EFFECTSTable 3-2 Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - Oral (continued) PBDEs70a Key to Species Figure (Strain) 66 Rat (CD) 67 Rat (Long-Evans) 68 Mouse (C57BL/6) 69 Mouse (C57BL/6N) 70 Mouse (NMRI) 71 Mouse (NMRI) Exposure/ Duration/ Frequency (Route) System NOAEL (mg/kg/day) Less Serious (mg/kg/day) LOAEL Serious (mg/kg/day) Reference Chemical Form Comments 10 d Gd 6-15 (GO) 25 50 (reduced fetal weight and increased skeletal variations associated with maternal tox) WIL Research Laboratories 1986 OctaBDE (DE-79) 14 d Gd 6-20 (GO) 1 10 (reduced serum T4 in fetuses) Zhou et al. 2002 PentaBDE (DE-71) once Pnd 10 (GO) 10 M Costa et al. 2015 TetraBDE (BDE47) No exposure-related changes in serum T3 or T4. once Pnd 10 (G) 6.8 M (decreased post-tetanic and long term potential in hippocampal slices at Pnd 17-19) Dingemans et al. 2007 TetraBDE (BDE47) once Pnd 10 (G) 0.8 M (altered spontaneous activity and habituation at 2-4 months) Eriksson et al. 2001 PentaBDE (BDE99) once Pnd 10 (G) 0.7 M 10.5 M (altered spontaneous activity and habituation at 2-4 months) Eriksson et al. 2001 TetraBDE (BDE47) 21898219082191821880.81221920.8219613. HEALTH EFFECTSTable 3-2 Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - Oral (continued) PBDEs71a Key to Species Figure (Strain) 72 Mouse (NMRI) 73 Mouse (NMRI) 74 Mouse (NMRI) 75 Mouse (NMRI) 76 Mouse (NMRI) 77 Mouse (C57BL/6N) Exposure/ Duration/ Frequency (Route) once Pnd 3 (G) System NOAEL (mg/kg) LOAEL Less Serious (mg/kg) Serious (mg/kg) 8 M (altered spontaneous activity and habituation at 4 months) Reference Chemical Form Eriksson et al. 2002b PentaBDE (BDE99) Comments once Pnd 10 (G) 8 M (altered spontaneous activity and habituation at 4 months) Eriksson et al. 2002b PentaBDE (BDE99) once Pnd 19 (G) 8 M Eriksson et al. 2002b PentaBDE (BDE99) No exposure-related alterations in spontaneous activity at 4 months. once Pnd 10 (G) 0.8 M 12 M (altered spontaneous activity and habituation at 4-6 months) Eriksson et al. 2006 PentaBDE (BDE99) once Pnd 10 (G) 0.8 M (decreased activity and impaired learning and memory during at 2-6 months) Fischer et al. 2008 PentaBDE (BDE99) once Pnd 10 (GO) 1 M (increased motor activity at 4 months) Gee and Moser 2008 TetraBDE (BDE47) 21973023731222430.40.82248823430.453. HEALTH EFFECTSTable 3-2 Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - Oral (continued) PBDEs72a Key to Species Figure (Strain) 78 Mouse (C57BL/6N) 79 Mouse (NMRI) 80 Mouse (C57BL/6) 81 Mouse (NMRI) 82 Mouse (NMRI) Exposure/ Duration/ Frequency (Route) System NOAEL (mg/kg/day) Less Serious (mg/kg/day) LOAEL Serious (mg/kg/day) Reference Chemical Form Comments once Pnd 10 (GO) 30 M Gee et al. 2008 TetraBDE (BDE47) No exposure-related changes in offspring body weight or serum T3 or T4 levels. once Pnd 10 (GO) 12 M (decreased spontaneous activity and impaired habituation at 2 months) Hallgren et al. 2015 PentaBDE (BDE99) once Pnd 10 (G) 0.4 0.8 (decreased total activity at 2 months) Sand et al. 2004 PentaBDE (BDE99) once Pnd 10 (G) 8 M (decreased spontaneous activity, altered habituation, and altered response to cholinergic agent at 2 months) Viberg et al. 2002 PentaBDE (BDE99) once Pnd 10 (G) 0.45 M (decreased spontaneous activity at 6 months) Viberg et al. 2003a HexaBDE (BDE153) 83 Mouse (C57/BL)Exposure/ Duration/ Frequency (Route) System NOAEL (mg/kg) LOAEL Reference Chemical Form Comments Less Serious Serious (mg/kg) (mg/kg) once Pnd 10 (G) 0.40.4 0.8 (decreased spontaneous activity and impaired habituation at 2-8 months) Viberg et al. 2004a PentaBDE (BDE99) 0.8once Pnd 10 (G) 0.40.4 M 12 M (decreased spontaneous activity, impaired habituation, and decreased density of cholinergic nicotinic receptors in hippocampus at 4 months) Viberg et al. 2004b PentaBDE (BDE99) 12once Pnd 3 (G) 16.8 M (decreased spontaneous activity and impaired habituation at 2 months) Viberg et al. 2006 OctaBDE (BDE203) 16.8once Pnd 10 (G) 16.8 M (decreased spontaneous activity, impaired habituation, and impaired learning and memory at 2-3 months) Viberg et al. 2006 OctaBDE (BDE203) 16.8once Pnd 3 (G) 15.2 M (decreased spontaneous activity at 2 months) Viberg et al. 2006 HeptaBDE (BDE183) 15.2224984 Mouse (NMRI) 225085 Mouse (NMRI) 232386 Mouse (NMRI) 232487 Mouse (NMRI) 23393. HEALTH EFFECTSTable 3-2 Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - Oral (continued) PBDEs73a Key to SpeciesFigure (Strain)234015.2234118.5234218.5204551523631103. HEALTH EFFECTSTable 3-2 Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - Oral (continued) Exposure/ LOAEL Duration/PBDEs74a FrequencyKey to Species NOAEL Less Serious Serious(Route)Figure (Strain) System (mg/kg) (mg/kg) (mg/kg) 88 Mouse once 15.2 MPnd 10(NMRI) (G) 89 Mouse once 18.5 MPnd 3(NMRI) (G) 90 Mouse once 18.5 M (decreased spontaneousPnd 10(NMRI) activity and impaired(G) habituation at 2 months) 91 Rabbit 13 d 5 F 15 F (delayed ossification ofGd 7-19(New sternebrae withZealand) (GO) decreased maternal weight gain) INTERMEDIATE EXPOSURE Systemic 92 Rat 5 wk Endocr 1 F 10 F (reduced serum T4Gd 6 - Pnd 21(Sprague- levels)Dawley) (IN) Reference Chemical Form Viberg et al. 2006 HeptaBDE (BDE183) Viberg et al. 2006 NonaBDE (BDE206) Viberg et al. 2006 NonaBDE (BDE206) Breslin et al. 1989 OctaBDE (technical) Bansal et al. 2014 PentaBDE (DE-71) Comments No change in spontaneous activity or habituation at 2 months. No changes in spontaneous motor behavior or habituation at 2 months. Dams were given DE-71-dosed vanilla wafers. 212433060212533306021400.55250.5 5252366330303. HEALTH EFFECTSTable 3-2 Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - Oral (continued) PBDEs75a Key to Figure 93 94 95 96 Species (Strain) Rat (Sprague-Dawley) Rat (Sprague-Dawley) Rat (Sprague-Dawley) Rat (Sprague-Dawley) Exposure/ Duration/ Frequency (Route) 15 d 1 x/d (GO) 15 d 1 x/d (GO) 21 wk Pmd 70 -Pnd 42 (GO) 6 wk Gd 1 - Pnd 21 (IN) System Endocr Bd Wt Hepatic Endocr Bd Wt Hepatic Renal Endocr Bd Wt Endocr Bd Wt NOAEL (mg/kg/day) 3 M 60 M 3 M 60 M 0.5 M 25 0.5 M 25 3 F 30 F LOAEL Less Serious Serious (mg/kg/day) (mg/kg/day) 30 M (follicular cell hypertrophy and hyperplasia in thyroid, reduced serum T3 and T4, increased serum TSH) 3 M (centrilobular hepatocyte hypertrophy) 30 M (decreased serum T3 and T4, increased serum TSH) 5 M (hepatocellular hypertrophy) 5 M (reduced serum T4) 30 F (reduced serum T3 and T4) Reference Chemical Form Becker et al. 2012 PentaBDE (DE-71) Becker et al. 2012 PentaBDE (DE-71) Bondy et al. 2011, 2013 PentaBDE (DE-71) Bowers et al. 2015 PentaBDE (DE-71) Comments Study 1 (conducted at ILS) Study 2 (conducted at RTI) Dams were given DE-71-dosed cookies. 21468821472 82821544020021554020021592 88403. HEALTH EFFECTSTable 3-2 Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - Oral (continued) a Key to Figure Species (Strain) Exposure/ Duration/ Frequency (Route) System NOAEL (mg/kg/day) Less Serious (mg/kg/day) LOAEL Serious (mg/kg/day) Reference Chemical Form Comments 97 Rat (Wistar) 21 d 1 x/d (GO) Hepatic Bd Wt 8 F (porphyria) 8 F (decreased body weight gain) Bruchajzer 2011 PentaBDE (technical) 98 Rat (Wistar) 28 d 1 x/d (GO) Hepatic Bd Wt 2 F 2 F 8 F (porphyria) 8 F (8% decrease in body weight gain) Bruchajzer 2011 PentaBDE (technical) 99 Rat (Wistar) 21 d 1 x/d Bd Wt 40 F 200 F (12% decrease in body weight gain) Bruchajzer et al. 2010 PentaBDE (technical) 100 Rat (Wistar) 28 d 1 x/d Bd Wt 40 F 200 F (14% decrease in body weight gain) Bruchajzer et al. 2010 PentaBDE (technical) 101 Rat (Wistar) 21 d 1 x/d (GO) Hepatic Bd Wt 2 F 8 F 8 F (porphyria) 40 F (9% decrease in body weight gain) Bruchajzer et al. 2012 OctaBDE (technical) PBDEs76216084020391.77204014.120372.4203819.221710.0150.015217417.526.23. HEALTH EFFECTSTable 3-2 Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - Oral (continued) a Key to Figure Species (Strain) Exposure/ Duration/ Frequency (Route) System NOAEL (mg/kg/day) Less Serious (mg/kg/day) LOAEL Serious (mg/kg/day) Reference Chemical Form Comments 102 Rat (Wistar) 28 d 1 x/d (GO) Bd Wt 8 F 40 F (8% decrease in body weight gain) Bruchajzer et al. 2012 OctaBDE (technical) 103 Rat (Sprague-Dawley) 90 d (GO) Hepatic 1.77 M Carlson 1980a PentaBDE (technical) 104 Rat (Sprague-Dawley) 90 d (GO) Hepatic 14.1 M Carlson 1980a PentaBDE (technical) 105 Rat (Sprague-Dawley) 90 d (GO) Hepatic 2.4 M Carlson, 1980a OctaBDE (technical) 106 Rat (Sprague-Dawley) 90 d (GO) Hepatic 19.2 M Carlson, 1980a OctaBDE (technical) 107 Rat (Sprague-Dawley) 90 d 1 x/d (GO) Hepatic Bd Wt 0.015 M 0.015 M Daubie et al. 2011 PentaBDE (BDE99) No exposure-related changes in hepatic clinical chemistry 108 Rat (Long- Evans) 125 d Pnd 1-125 (F) Endocr Bd Wt 26.2 M 17.5 M (reduced serum T4) Driscoll et al. 2009 PentaBDE (DE-71) PBDEs772177505050502185181818234620220202203. HEALTH EFFECTSTable 3-2 Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - Oral (continued) PBDEs78a Key to Species Figure (Strain) 109 Rat (Wistar) 110 Rat (Sprague-Dawley) 111 Rat (Sprague-Dawley) Exposure/ Duration/ Frequency (Route) 18 wk Gd 6 - Pnw 16 5-7 d/wk (GO) 33 d Gd 6 - Pnd 18 (GO) 70 d (F) LOAEL System NOAEL (mg/kg/day) Less Serious (mg/kg/day) Serious (mg/kg/day) Reference Chemical Form Comments Hepatic 50 (hepatocellular hypertrophy and vacuolization) Dunnick et al. 2012 PentaBDE (DE-71) Endocr 50 (thyroid gland follicular hypertrophy) Bd Wt 50 M 50 F (14% decrease in body weight) Hepatic 18 F (reduced liver vitamin A) Ellis-Hutchings et al. 2006 PentaBDE (DE-71) Endocr Bd Wt 18 F 18 F (reduced serum T4) Hepatic Endocr 20 M 2 M 20 M (reduced serum T4; increased epithelial thickness of inner follicles and vacuolation of the luminal apices of epithelial cells in thyroid) Ernest et al. 2012 52.1% penta-, 44.2% deca-, 0.4% octa-BDE No exposure-related changes in hepatic clinical chemistry. Bd Wt 20 M Metab 2 M 20 M (reduced serum glucose level) 23592.525220614141420539909090205599090903. HEALTH EFFECTSTable 3-2 Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - Oral (continued) PBDEs79a Key to Species Figure (Strain) 112 Rat (Sprague-Dawley) 28 d (G) 113 Rat (Sprague-Dawley) 4 wk 1 x/d (GO) 114 Rat (CD) 28 d (F) 115 Rat (CD) 28 d (F) Exposure/ Duration/ Frequency (Route) System Hepatic Endocr Bd Wt Metab Hepatic Renal Endocr Bd Wt Hepatic Renal Endocr Bd Wt LOAEL NOAEL (mg/kg/day) Less Serious (mg/kg/day) Serious (mg/kg/day) Reference Chemical Form Comments 2.5 25 (decreased hepatic vitamin A content) Fattore et al. 2001 PentaBDE (Bromkal 70-5DE) 14 M (reduced serum T4) Hoppe and Carey 2007 PentaBDE (technical) 14 M 14 M (increased adipocyte lipolysis, decreased adipocyte glucose oxidation) 9 (increased liver weight and enlarged parenchymal cells) IRDC 1976 PentaBDE (technical) 90 90 90 9 (increased liver weight and enlarged parenchymal cells) IRDC 1976 OctaBDE (technical) 90 90 90 2077750750750707505506007507507507060022091.7 10.230.63. HEALTH EFFECTSTable 3-2 Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - Oral (continued) Exposure/ Duration/a Key to Species Frequency Figure (Strain) (Route) PBDEs80116 Rat 13 wk (CD) (F) 117 Rat 36 d Gd 6 - Pnd 21(Long- Evans) (GO) System Resp Cardio Gastro Hemato Hepatic Renal Endocr Dermal Ocular Bd Wt Endocr Bd Wt LOAEL NOAEL (mg/kg/day) Less Serious (mg/kg/day) Serious (mg/kg/day) Reference Chemical Form Comments 750 F 750 F IRDC 1977 OctaBDE (technical) 750 F 70 F 750 F (reduced erythrocytes, hematocrit and hemoglobin) 5 M (cytomegaly with vacuolation and necrosis at higher doses) 50 M 600 M (minimal increase in tubular degenerative changes) 7 F 50 M (increased thyroid weight with follicular epithelial changes at higher doses) 750 F 750 F 70 F 600 M (12% reduced body weight gain) 1.7 F 10.2 F (reduced serum T4) Kodavanti et al. 2010 PentaBDE (DE-71) 30.6 F 22622502502502.52525025025250226111.422923 30603. HEALTH EFFECTSTable 3-2 Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - Oral (continued) Exposure/ Duration/a Key to Species Frequency Figure (Strain) (Route) PBDEs81118 Rat 28 d 1 x/d(Sprague-Dawley) (GO) 119 Rat ~11 wk Pmd 28 -(Long- Evans) Pnd 21 (IN) 120 Rat 20 d Pnd 22-41(Wistar) (GO) System Resp Cardio Hemato Hepatic Renal Bd Wt Metab Bd Wt Endocr Bd Wt LOAEL NOAEL (mg/kg/day) Less Serious (mg/kg/day) Serious (mg/kg/day) Reference Chemical Form Comments 250 Oberg et al. 2010 PentaBDE (Bromkal 70-5DE) 250 250 2.5 25 (centrilobular hypertrophy, reduced vitamin A content in liver) 250 250 25 250 (hypercalcemia, magnesemia, and phosphatemia in males; hyperatremia and hypokalemia in females) 11.4 F Poon et al. 2011 PentaBDE (DE-71) Rats were given pentaBDE-dosed vanilla wafers. 3 F 30 F (reduced serum T4) Stoker et al. 2004 PentaBDE (DE-71) 60 F 2295360230160120 3. HEALTH EFFECTSTable 3-2 Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - Oral (continued) a Key to Figure Species (Strain) Exposure/ Duration/ Frequency (Route) System NOAEL (mg/kg/day) Less Serious (mg/kg/day) LOAEL Serious (mg/kg/day) Reference Chemical Form Comments 121 Rat (Wistar) 31 d Pnd 23-53 (GO) Endocr Bd Wt 60 M 3 M (reduced serum T4) Stoker et al. 2004 PentaBDE (DE-71) 122 Rat (Wistar) 31 d Pnd 23-53 (GO) Endocr Bd Wt 120 M 60 M (reduced serum T4) Stoker et al. 2005 PentaBDE (DE-71) PBDEs8223152002002002000.052001.1200 9.7 66.722753.23. HEALTH EFFECTSTable 3-2 Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - Oral (continued) Exposure/ Duration/a Key to Species Frequency Figure (Strain) (Route) 123 Rat (Wistar) 28 d (G) 124 Rat (Wistar) 34 d Gd 1 - Pnd 14 (F) PBDEs83System Resp Cardio Gastro Musc/skel Hepatic Renal Endocr Dermal Bd Wt Metab Endocr LOAEL NOAEL (mg/kg/day) Less Serious (mg/kg/day) Serious (mg/kg/day) Reference Chemical Form Comments 200 Van der Ven et al. 2008b PentaBDE (DE-71) NOAEL values for Bd Wt, Hepatic, Endocr, and Metab effects are BMDL(RD10%) values for decreased Bd Wt, vitamin A in liver, serum T4 and glucose. 200 200 200 0.05 M 200 1.1 M 200 9.7 M66.7 M 3.2 F (reduced serum T4) Wang et al. 2011a TetraBDE (BDE47) 125 Rat Table 3-2 Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - Oral (continued) xposure/ LOAEL uration/ requency NOAEL Less Serious Serious (Route) System (mg/kg/day) (mg/kg/day) (mg/kg/day) Reference Chemical Form Comments 90 d Resp 100 (F) 100WIL Research Laboratories 1984 PentaBDE (DE-71) Cardio 100 100Gastro 100 100Hemato 100 100Musc/skel 100 100Hepatic 2 (hypertrophy, mild degeneration, and slight necrosis) 2Renal 100 100Endocr 2 10 (reduced serum T4) 2 10Dermal 100 100Ocular 100 100Bd Wt 10 100 (reduced weight gain) 10 10030 d Bd Wt 1 M 1 x/d 1(GO) Yan et al. 2012 TetraBDE (BDE47) 6 wk Bd Wt 0.2 F Gd 1 - Pnd 21 0.2Zhao et al. 2014 (GO) PentaBDE (BDE99) (Sprague-Dawley) 2048126 Rat (Sprague-Dawley) 2361127 Rat (Sprague-Dawley) 23833. HEALTH EFFECTSPBDEs84a Key to Species Figure (Strain) EDF20871030302143182212122290.450.450.453. HEALTH EFFECTSTable 3-2 Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - Oral (continued) PBDEs85a Key to Figure 128 129 130 131 Species (Strain) Rat (Long-Evans) Mouse (CD-1) Mouse (C57BL/6J) Mouse (BALB/c) Exposure/ Duration/ Frequency (Route) 36 d Gd 6 - Pnd 21 (GO) 36 d Gd 6 - Pnd 21 70-80 d Pmd 28 -Pnd 21 (IN) 28 d (F) System Endocr Bd Wt Bd Wt Bd Wt Hepatic Endocr Bd Wt NOAEL (mg/kg/day) 10 F 30 F 18 F 1 F 0.45 F LOAEL Less Serious Serious (mg/kg/day) (mg/kg/day) 30 F (reduced maternal serum T4) 0.45 F (hepatocyte vacuolation, pyknotic nuclei in the hepatocytes, periportal lymphocytic infiltration) 0.45 F (cellular debris in the follicular lumen of thyroid; increased serum testosterone and E2) Reference Chemical Form Zhou et al. 2002 PentaBDE (DE-71) Branchi et al. 2005 PentaBDE (BDE99) Koenig et al. 2012 TetraBDE (BDE47) Maranghi et al. 2013 TetraBDE (BDE47) Comments Administered via "self-administration" from a modified syringe or gavage. Mice were fed 1-2 tetraBDE dosed cornflakes. 237611122684524522271452452228912279303. HEALTH EFFECTSTable 3-2 Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - Oral (continued) PBDEs86a Key to Species Figure (Strain) 132 Mouse (NS) 6 wk 5 d/wk (GO) 133 Mouse (NMRI) 29 d Gd 4 - Pnd 17 ~every 3 d (GO) 134 Mouse (NMRI) 29 d Gd 4 - Pnd 17 ~every 3 d (GO) 135 Mouse (C57BL/6J) 70-80 d Pmd 28 -Pnd 21 (IN) 136 Mouse (C56BL/6) 30 d (GO) Exposure/ Duration/ Frequency (Route) System Hepatic Bd Wt Metab Endocr Bd Wt Endocr Bd Wt Bd Wt Bd Wt LOAEL NOAEL (mg/kg/day) Less Serious (mg/kg/day) Serious (mg/kg/day) Reference Chemical Form 1 M McIntyre et al. 2015 TetraBDE (BDE47) 1 M 1 M 452 F Skarman et al. 2005 PentaBDE (BDE99) 452 F 452 F Skarman et al. 2005 PentaBDE (Bromkal 70-5DE) 452 F 1 F Ta et al. 2011 TetraBDE (BDE47) 30 M Wang et al. 2013 TetraBDE (BDE47) Comments No exposure-related changes in glucose tolerance, insulin resistance, lipogenesis, or liver weight or histology. BDE-99; No change in maternal serum T4 levels. Bromkal 70-5DE; no change in maternal serum T4 levels. Dams were fed tetraBDE-dosed cornflakes. Table 3-2 Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - Oral (continued) a Key to Figure Species (Strain) Exposure/ Duration/ Frequency (Route) System NOAEL (mg/kg/day) LOAEL Comments Less Serious (mg/kg/day) Serious (mg/kg/day) Reference Chemical Form 137 138 139 140 Mouse (ICR) 2117Mouse (ICR) 2379Mouse (ICR) 2380Mink (NS) 223128 d 1 x/d (GO) 12 wk 7 d/wk (GO) 12 wk 7 d/wk (GO) 9 wk (F) Hepatic Renal Bd Wt Hepatic Bd Wt Hepatic Bd Wt Bd Wt 1.21501500.081.2 M 150 M 150 M 0.08 M 1.21.21501500.631.2 M (swollen hepatic cells) 1.2 M (10% decrease in relative kidney weight) 150 M (increased relative liver weight, increased serum ALT, hepatocyte hypertrophy and vacuolization and inflammatory cell infiltration) 150 M (increased relative liver weight, increased serum ALT, hepatocyte hypertrophy and vacuolization and inflammatory cell infiltration) 0.63 M (21% decrease in body weight) Zhang et al. 2014 DiBDE (BDE15) Zhang et al. 2015b TetraBDE (BDE47) Zhang et al. 2015a TetraBDE (BDE47) Martin et al. 2007 PentaBDE (DE-71) 3. HEALTH EFFECTSPBDEs8723330.3123380.062138252347902350902351903. HEALTH EFFECTSTable 3-2 Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - Oral (continued) PBDEs88Exposure/ LOAEL Duration/a FrequencyKey to Species NOAEL Less Serious Serious(Route)Figure (Strain) 141 Mink (NS) 142 Mink (NS) Immuno/ Lymphoret 143 Rat (Sprague-Dawley) 144 Rat (CD) 145 Rat (CD) 146 Rat (CD) System (mg/kg/day) (mg/kg/day) (mg/kg/day) 16-17 wk Bd Wt 0.31 FPmw 4 -Pnw 6 (F) 43-44 wk Bd Wt 0.06Pmw 4 -Pnw 33 (F) 21 wk 25Pmd 70 -Pnd 42 (GO) 28 d 90(F) 28 d 90(F) 28 d 90(F) Reference Chemical Form Zhang et al. 2009 PentaBDE (DE-71) Zhang et al. 2009 PentaBDE (DE-71) Bondy et al. 2011, 2013 PentaBDE (DE-71) IRDC 1976 PentaBDE (technical) IRDC 1976 OctaBDE (technical) IRDC 1976 PentaBDE (technical) Comments No exposure-related changes in spleen or thymus weight or histology. No exposure-related changes in spleen weight or spleen or thymus histology. No exposure-related changes in spleen weight or spleen or thymus histology. No exposure-related changes in spleen weight or spleen or thymus histology. 2078Table 3-2 Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - Oral (continued) es n) Exposure/ LOAEL Reference Chemical Form Duration/ Frequency NOAEL Less Serious Serious (Route) System (mg/kg/day) (mg/kg/day) (mg/kg/day) 13 wk 750 F (F) 750IRDC 1977 OctaBDE (technical) ue-) 28 d 250 1 x/d 250(GO) Oberg et al. 2010 PentaBDE (Bromkal 70-5DE)ue-90 d 100 (F) 100WIL Research Laboratories 1984 ) PentaBDE (DE-71) /c) 28 d 0.45 F (follicular hyperplasia and (F) lymphocytic infiltration in spleen; lymphocytic Maranghi et al. 2013 TetraBDE (BDE47) apoptosis and Hassal's bodies in thymus) 0.459 wk 0.08 M 0.63 M (spleen hyperplasia) (F) 0.08 0.63Martin et al. 2007 PentaBDE (DE-71) 22562049222722303. HEALTH EFFECTSPBDEs89a Key to SpeciFigure (Strai147 Rat (CD) 148 Rat (SpragDawley149 Rat (SpragDawley150 Mouse(BALB151 Mink (NS) Comments No exposure-related changes in spleen weight or spleen or thymus histology. No exposure-related changes in spleen or thymus weight or histology. No exposure-related changes in thymus weight or spleen or thymus histology. Immune function was not altered (KLH antibody induction, PHA skin challenge). 21700.015217317.526.22257250231320022550.121630.253. HEALTH EFFECTSTable 3-2 Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - Oral (continued) PBDEs90Exposure/ LOAEL Duration/a FrequencyKey to Species NOAEL Less Serious Serious(Route)Figure (Strain) System (mg/kg/day) (mg/kg/day) (mg/kg/day) Neurological 152 Rat 90 d 0.015 M1 x/d(Sprague-Dawley) (GO) 153 Rat 125 d 17.5 M 26.2 M (impaired attention andPnd 1-125(Long- Evans) inhibitory control)(F) 154 Rat 28 d 2501 x/d(Sprague-Dawley) (GO) 155 Rat 28 d 200(Wistar) (G) 156 Rat 30 d 0.1 M (impaired learning and1 x/d(Sprague- memory; decreasedDawley) (GO) glutamate and receptor density in hippocampus) 157 Mink 19 wk 0.25 FPmd 28 -(NS) Pnw 6 (F) Reference Chemical Form Daubie et al. 2011 PentaBDE (BDE99) Driscoll et al. 2009 PentaBDE (DE-71) Oberg et al. 2010 PentaBDE (Bromkal 70-5DE) Van der Ven et al. 2008b PentaBDE (DE-71) Yan et al. 2012 TetraBDE (BDE47) Bull et al. 2007 PentaBDE (DE-71) Comments No exposure-related changes in elevated plus-maze, open-field, or Morris water maze test. No exposure-related changes in brain weight or histology. No exposure-related changes in brain weight or histology. No maternal cholinergic effects were observed. 212230602123602139252365303. HEALTH EFFECTSTable 3-2 Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - Oral (continued) PBDEs91a Key to Species Figure (Strain) Reproductive 158 Rat (Sprague-Dawley) 159 Rat (Sprague-Dawley) 160 Rat (Sprague-Dawley) 161 Rat (Sprague-Dawley) Exposure/ Duration/ Frequency (Route) System NOAEL (mg/kg/day) Less Serious (mg/kg/day) LOAEL Serious (mg/kg/day) Reference Chemical Form 15 d 1 x/d (GO) 30 M 60 M (67% increase in serum prolactin; dose-related increase in serum testosterone and FSH) Becker et al. 2012 PentaBDE (DE-71) 15 d 1 x/d (GO) 60 M Becker et al. 2012 PentaBDE (DE-71) 21 wk Pmd 70 -Pnd 42 (GO) 25 Bondy et al. 2011, 2013 PentaBDE (DE-71) 6 wk Gd 1 - Pnd 21 (IN) 30 F Bowers et al. 2015 PentaBDE (DE-71) Comments Study 1 (conducted at ILS) Study 2 (conducted at RTI); no dose-related changes in reproductive organ weight, histopathology, or serum reproductive hormone levels. No treatment-related changes in the number of pregnant females, litters, or litter size. Dams were given DE-71-dosed cookies. No treatment-related changes in no. of implants, no. of litters, litter size, or sex ratio. 2166Table 3-2 Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - Oral (continued) Exposure/ LOAEL Reference Chemical Form Duration/ Frequency NOAEL Less Serious Serious (Route) System (mg/kg/day) (mg/kg/day) (mg/kg/day) 36 d 2 F Gd 6 - Pnd 21 2Cheng et al. 2009 (GO) PentaBDE (BDE99) 70 d 20 M (F) 20Ernest et al. 2012 52.1% penta-, 44.2% dec0.4% octa-BDE 8 wk 0.001 M 0.03 M (increased epithelial (GO) 0.001 thickness in testes, spermatocyte apoptosis) Huang et al. 2015 TetraBDE (BDE47) 0.0328 d 90 (F) 90IRDC 1976 PentaBDE (technical) 28 d 90 (F) 90IRDC 1976 OctaBDE (technical) 23452374234823523. HEALTH EFFECTSPBDEs92a Key to Species Figure (Strain) 162 Rat (Sprague-Dawley) 163 Rat (Sprague-Dawley) 164 Rat (Sprague-Dawley) 165 Rat (CD) 166 Rat (CD) a-, Comments No exposure-related changes in number of litters, litter size, or sex ratio. No exposure-related changes in reproductive organ weight, testicular histology, or sperm parameters. No exposure-related changes in reproductive organ weight or histology. No exposure-related changes in reproductive organ weight or histology. 20806007502258250226011.42291306022943303. HEALTH EFFECTSTable 3-2 Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - Oral (continued) PBDEs93Exposure/ LOAEL Duration/a FrequencyKey to Species NOAEL Less Serious Serious(Route)Figure (Strain) System (mg/kg/day) (mg/kg/day) (mg/kg/day) 167 Rat 13 wk 600 M(CD) (F) 750 F 168 Rat 28 d 2501 x/d(Sprague-Dawley) (GO) 169 Rat ~11 wk 11.4 FPmd 28 -(Long- Evans) Pnd 21 (IN) 170 Rat 20 d 30 F 60 F (delayed vaginalPnd 22-41(Wistar) opening)(GO) 171 Rat 31 d 3 M 30 M (delayed preputialPnd 23-53(Wistar) separation)(GO) Reference Chemical Form IRDC 1977 OctaBDE (technical) Oberg et al. 2010 PentaBDE (Bromkal 70-5DE) Poon et al. 2011 PentaBDE (DE-71) Stoker et al. 2004 PentaBDE (DE-71) Stoker et al. 2004 PentaBDE (DE-71) Comments No exposure-related changes in reproductive organ weight or histology No exposure-related changes in reproductive organ weight or histology. Rats were given pentaBDE-dosed vanilla wafers; no change in number of pregnancies, implantation sites, or live pups, litter size, or sex ratio. 2300602314 9.6200205110022800.00123820.23. HEALTH EFFECTSTable 3-2 Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - Oral (continued) PBDEs94a Key to Species Figure (Strain) 172 Rat (Wistar) 173 Rat (Wistar) 174 Rat (Sprague-Dawley) 175 Rat (Sprague-Dawley) 176 Rat (Sprague-Dawley) Exposure/ Duration/ Frequency (Route) System NOAEL (mg/kg/day) Less Serious (mg/kg/day) LOAEL Serious (mg/kg/day) Reference Chemical Form 31 d Pnd 23-53 (GO) 60 M (delayed PPS, 22-28% decrease in prostate and seminal vesicle weights) Stoker et al. 2005 PentaBDE (DE-71) 28 d (G) 9.6 M Van der Ven et al. 2008b PentaBDE (DE-71) 200 F 90 d (F) 100 WIL Research Laboratories 1984 PentaBDE (DE-71) 8 wk 6 d/wk (GO) c 0.001 M (34% reduction in serum testosterone) Zhang et al. 2013b TetraBDE (BDE47) 6 wk Gd 1 - Pnd 21 (GO) 0.2 M Zhao et al. 2014 PentaBDE (BDE99) Comments Male NOAEL is a BMDL(RD10%) for increased % of deformed sperm heads; no exposure-related changes in female reproductive organ weights or histology. No exposure-related changes in reproductive organ weight or histology. No exposure-related changes in gestational lengths, litter sizes, sex ratio, or live births. 2112302142182211122280.452234503. HEALTH EFFECTSTable 3-2 Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - Oral (continued) PBDEs95a Key to Species Figure (Strain) 177 Rat (Long-Evans) 178 Mouse (CD-1) 179 Mouse (C57BL/6J) 180 Mouse (BALB/c) 181 Mouse (BALB/c) Exposure/ Duration/ Frequency (Route) System NOAEL (mg/kg/day) Less Serious (mg/kg/day) LOAEL Serious (mg/kg/day) Reference Chemical Form 36 d Gd 6 - Pnd 21 (GO) 30 F Zhou et al. 2002 PentaBDE (DE-71) 36 d Gd 6 - Pnd 21 18 F Branchi et al. 2005 PentaBDE (BDE99) 70-80 d Pmd 28 -Pnd 21 (IN) 1 F Koenig et al. 2012 TetraBDE (BDE47) 28 d (F) 0.45 F (increased serum testosterone and E2) Maranghi et al. 2013 TetraBDE (BDE47) 34 d 1 x/d (GO) 50 F Mercado-Feliciano and Bigsby 2008a PentaBDE (DE-71) Comments No change in the gestation length, litter size, or sex ratio Administered via "self-administration" from a modified syringe or gavage; No change in gestation length, litter size, # live pups, or sex ratio. Mice were fed 1-2 tetraBDE dosed cornflakes; no change in the number of pregnancies, size of litter, or sex ratio of pups. Mice were ovariectomized 3 weeks prior to exposure; no change in uterus weight or uterus/vaginal histology. 226745222704522288122780.00150.04523540.030.13. HEALTH EFFECTSTable 3-2 Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - Oral (continued) PBDEs96a Key to Species Figure (Strain) 182 Mouse (NMRI) 183 Mouse (NMRI) 184 Mouse (C57BL/6J) 185 Mouse (C56BL/6) 186 Mouse (C57BL/6) Exposure/ Duration/ Frequency (Route) System NOAEL (mg/kg/day) Less Serious (mg/kg/day) LOAEL Serious (mg/kg/day) Reference Chemical Form 29 d Gd 4 - Pnd 17 ~every 3 d (GO) 452 F Skarman et al. 2005 PentaBDE (BDE99) 29 d Gd 4 - Pnd 17 ~every 3 d (GO) 452 F Skarman et al. 2005 PentaBDE (Bromkal 70-5DE) 70-80 d Pmd 28 -Pnd 21 (IN) 1 F Ta et al. 2011 TetraBDE (BDE47) 30 d (GO) 0.0015 M 0.045 M (germ cell loss and apoptosis in testes) Wang et al. 2013 TetraBDE (BDE47) 10 wk Pmd 28 -Pnd 21 (GO) 0.03 F 0.1 F (58% decrease in litters surviving until Pnd 8) Woods et al. 2012 TetraBDE (BDE47) Comments BDE-99; no change in maternal serum T4 levels. Bromkal 70-5DE; no changes in pregnancy rate, gestation length, or litter size. Dams were fed tetraBDE-dosed cornflakes; no change in gestation length, litter size, or sex ratio. Females carrying on copy of a truncated Mecp2 gene were mated to unexposed wild-type males; LSE values are based on the wild-type offspring only. 21640.05 0.2523320.06 0.31236211021331221370.553. HEALTH EFFECTSTable 3-2 Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - Oral (continued) PBDEs97a Key to Species Figure (Strain) 187 Mink (NS) 188 Mink (NS) Developmental 189 Rat (Sprague-Dawley) 190 Rat (Sprague-Dawley) 191 Rat (Sprague-Dawley) Exposure/ Duration/ Frequency (Route) System NOAEL (mg/kg/day) Less Serious (mg/kg/day) LOAEL Serious (mg/kg/day) Reference Chemical Form Comments 19 wk Pmd 28 -Pnw 6 (F) 0.05 F 0.25 F (no litters produced) Bull et al. 2007 PentaBDE (DE-71) 16-17 wk Pmw 4 -Pnw 6 (F) 0.06 F 0.31 F (complete litter loss) Zhang et al. 2009 PentaBDE (DE-71) 5 wk Gd 6 - Pnd 21 (IN) 1 M 10 M (reduced serum T4 in Pnd 21 offspring) Bansal et al. 2014 PentaBDE (DE-71) Dams were given DE-71-dosed vanilla wafers. 36 d Gd 6 - Pnd 21 (GO) 1 2 (altered neurobehavior, decreased hippocampal BDNF, and decreased serum T3, T4, and free T4 in offspring at Pnd 21-23) Blanco et al. 2013 PentaBDE (BDE99) 21 wk Pmd 70 -Pnd 42 (GO) 0.5 M 5 M (decreased serum T4 in male offspring on Pnd 43) Bondy et al. 2011, 2013 PentaBDE (DE-71) 23640.332165221751521821822081.710.23. HEALTH EFFECTSTable 3-2 Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - Oral (continued) a Key to Figure Species (Strain) Exposure/ Duration/ Frequency (Route) System LOAEL NOAEL (mg/kg/day) Less Serious (mg/kg/day) Serious (mg/kg/day) Reference Chemical Form Comments 192 Rat (Sprague-Dawley) 6 wk Gd 1 - Pnd 21 (IN) 0.3 3 (reduced serum T3 and T4 in Pnd 21 offspring) Bowers et al. 2015 PentaBDE (DE-71) Dams were given DE-71-dosed cookies. 193 Rat (Sprague-Dawley) 36 d Gd 6 - Pnd 21 (GO) 2 M (delayed appearance of reflexes, impaired learning/memory at Pnd 36-37, and oxidative stress in the hippocampus at Pnd 37) Cheng et al. 2009 PentaBDE (BDE99) 194 Rat (Long- Evans) 7 d Pnd 6-12 (GO) 15 M Driscoll et al. 2012 PentaBDE (DE-71) No learning or attention deficits at Pnd 40-95. 195 Rat (Sprague-Dawley) 33 d Gd 6 - Pnd 18 (GO) 18 (decreased serum T4 in offspring) Ellis-Hutchings et al. 2006 PentaBDE (DE-71) Half of the dams in each group were maintained on a vitamin A deficient diet. 196 Rat (Long- Evans) 36 d Gd 6 - Pnd 21 (GO) 1.7 10.2 (reduced serum T4 in offspring at Pnd 7 and 14, reduced mammary gland development at Pnd 21, reduced female body weight from Pnd 29-58) Kodavanti et al. 2010 PentaBDE (DE-71) PBDEs9822360.962.85223711.222595.722441.722451.710.222743.23. HEALTH EFFECTSTable 3-2 Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - Oral (continued) a Key to Figure Species (Strain) Exposure/ Duration/ Frequency (Route) System NOAEL (mg/kg/day) Less Serious (mg/kg/day) LOAEL Serious (mg/kg/day) Reference Chemical Form Comments 197 Rat (Long- Evans) 36 d Gd 6 - Pnd 21 (GO) 0.96 M 2.85 M (reduced serum T4 in male offspring at Pnd 7-21) Miller et al. 2012 PentaBDE (DE-71) 198 Rat (Long- Evans) 36 d Gd 6 - Pnd 21 (GO) 11.2 (reduced serum T4 in offspring at Pnd 7-21) Miller et al. 2012 PentaBDE (DE-71) 199 200 Rat (Long- Evans) ~11 wk Pmd 28 -Pnd 21 (IN) Rat (Long- Evans) 36 d Gd 6 - Pnd 21 (GO) 5.7 (reduced serum T4) 1.7 M (hypertensive reaction to hyperosmotic stress in adult male offspring ) Poon et al. 2011 PentaBDE (DE-71) Shah et al. 2011 PentaBDE (DE-71) Rats were given pentaBDE-dosed vanilla wafers. 201 Rat (Long- Evans) 36 d Gd 6 - Pnd 21 (GO) 1.7 10.2 (transient reduction in serum T4 in offspring at Pnd 4 and 21) Szabo et al. 2009 PentaBDE (DE-71) 202 Rat (Wistar) 34 d Gd 1 - Pnd 14 (F) 3.2 (reduced serum T4 in offspring at Pnd 7 and 14) Wang et al. 2011a TetraBDE (BDE47) PBDEs9923810.2208611021411823750.222100.033. HEALTH EFFECTSTable 3-2 Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - Oral (continued) PBDEs100a Key to Species Figure (Strain) 203 Rat (Sprague-Dawley) 204 Rat (Long-Evans) 205 Mouse (CD-1) 206 Mouse (CD-1) 207 Mouse (C57BL/6J) Exposure/ Duration/ Frequency (Route) System NOAEL (mg/kg/day) Less Serious (mg/kg/day) LOAEL Serious (mg/kg/day) Reference Chemical Form 6 wk Gd 1 - Pnd 21 (GO) 0.2 M Zhao et al. 2014 PentaBDE (BDE99) 36 d Gd 6 - Pnd 21 (GO) 1 10 (reduced serum T4 in offspring on Pnd 4 and 14) Zhou et al. 2002 PentaBDE (DE-71) 36 d Gd 6 - Pnd 21 18 M (transient alterations in open-field behavior of offspring at Pnd 34) Branchi et al. 2005 PentaBDE (BDE99) 5 wk Gd 8 - Pnd 21 (GO) 0.2 M Kim et al. 2015 TetraBDE (BDE47) 70-80 d Pmd 28 -Pnd 21 (IN) 0.03 (impaired learning in offspring at Pnw 8) Koenig et al. 2012 TetraBDE (BDE47) Comments No exposure-related changes in reflex maturation, motor coordination, or spatial learning of Pnd 3-36 offspring; no offspring body weight effects. Administered via "self-administration" from a modified syringe or gavage; data from both groups were pooled for neurodevelopmental endpoints. No exposure-related changes in offspring body weight, motor activity at Pnd 21, or social interaction at Pnd 70. Mice were fed 1-2 tetraBDE dosed cornflakes. 2266452226945222870.0322540.0321610.053. HEALTH EFFECTSTable 3-2 Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - Oral (continued) PBDEs101a Key to Species Figure (Strain) 208 Mouse (NMRI) 209 Mouse (NMRI) 210 Mouse (C57BL/6J) 211 Mouse (C57BL/6) 212 Mink (NS) Exposure/ Duration/ Frequency (Route) System NOAEL (mg/kg/day) Less Serious (mg/kg/day) LOAEL Serious (mg/kg/day) Reference Chemical Form 29 d Gd 4 - Pnd 17 ~every 3 d (GO) 452 Skarman et al. 2005 PentaBDE (BDE99) 29 d Gd 4 - Pnd 17 ~every 3 d (GO) 452 (reduced serum T4 in offspring at Pnd 11) Skarman et al. 2005 PentaBDE (Bromkal 70-5DE) 70-80 d Pmd 28 -Pnd 21 (IN) 0.03 F (decreased center-field activity in open field in female offspring at Pnd 60) Ta et al. 2011 TetraBDE (BDE47) 10 wk Pmd 28 -Pnd 21 (GO) 0.03 F (decreased pre-weaning weight; decreased pup vocalizations on Pnd 8-10, decreased sociability on Pnd 72) Woods et al. 2012 TetraBDE (BDE47) 19 wk Pmd 28 -Pnw 6 (F) 0.05 Bull et al. 2007 PentaBDE (DE-71) Comments BDE-99; no changes in offspring body weight or serum T4 levels Bromkal 70-5DE Dams were fed tetraBDE-dosed cornflakes. Females carrying on copy of a truncated Mecp2 gene were mated to unexposed wild-type males; LSE values are based on the wild-type offspring only. No cholinergic effects in 6-week-old offspring. Table 3-2 Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - Oral (continued) Exposure/ LOAEL a Key to Figure Species (Strain) Duration/ Frequency (Route) System NOAEL (mg/kg/day) Reference Chemical Form Comments Less Serious (mg/kg/day) Serious (mg/kg/day) 213 Mink (NS) 40 wk Pmd 28 -Pnw 27 (F) 0.050.05 Bull et al. 2007 PentaBDE (DE-71) No cholinergic effects in 45-week-old offspring. 2162214 Mink (NS) 16-17 wk Pmw 4 -Pnw 6 0.060.06 Zhang et al. 2009 PentaBDE (DE-71) No change in body weight, organ weights, plasma T3/T4, hepatic (F) enzyme activity, or thyroid histology in offspring at weaning (Pnw 6). 23343. HEALTH EFFECTSPBDEs102a The number corresponds to entries in Figure 3-2. b Three studies were used to derive an acute oral minimal risk level (MRL); concentration divided by an uncertainty factor of 1000 (10 for use of a LOAEL, 10 for animal to human, and 10 for human variability), resulting in an MRL of 0.00006 mg/kg/day. c Used to derive an intermediate oral minimal risk level (MRL); concentration divided by an uncertainty factor of 300 (3 for use of a minimal LOAEL, 10 for animal to human, 10 for human variability), resulting in an MRL of 0.000003 mg/kg/day. Note on chemical form: Mixtures are identified by composition or trade name (if reported); otherwise, they are reported as "technical". Trade names include Bromkal 70, Bromakal 70-5 DE, DE-71, and Saytex 115 for pentaBDE mixtures and DE-79, FR-1208, and Saytex 111 for octaBDE mixtures. For the studies by Bruchajzer (2011) and Bruchajzer et al. (2010, 2011, 2012), the mixtures were made to resemble formerly used commercial mixtures. The pentaBDE mixture was composed of 63.2% pentaBDE, 21.4% tetraBDE, 15.4% hexaBDE, and 0.04% heptaBDE; the octaBDE mixture was composed of 65.7% octaBDE, 14.8% heptaBDE, 1.7% hexaBDE, and 17.8% nona- and deca-BDE. Individual congeners are identified by IUPAC number: BDE 15 = 4,4'-diBDE; BDE 47 = 2,2',4,4'-tetraBDE; BDE 99 = 2,2',4,4',5-pentaBDE; BDE 153 = 2,2',4,4',5,5'-hexaBDE; BDE 183 = 2,2',3,4,4',5',6-heptaBDE; BDE 203 = 2,2',3,4,4',5,5',6-octaBDE; BDE 206 = 2,2',3,3',4,4',5,5',6-nonaBDE BDNF = brain derived neurotrophic factor; Bd Wt = body weight; BMDL = benchmark dose lower confidence limit; Cardio = cardiovascular; d = day(s); E2 = estradiol; Endocr = endocrine; (F) = feed; F = Female; FSH = follicle stimulating hormone; (G) = gavage; Gastro = gastrointestinal; Gd = gestational day; (GO) = gavage in oil; Hemato = hematological; IgG = immunoglobulin G; Immuno/Lymphoret = immunological/lymphoreticular; IN = ingestion; LH = luteinizing hormone; KLH = keyhole limpet hemocyanin; LD50 = lethal dose, 50% kill; LOAEL = lowest-observed-adverse-effect level; M = male; Metab = metabolism; Musc/skel = musculoskeletal; NOAEL = no-observed-adverse-effect level; NS = not specified; PHA = phytohemagglutinin; Pmd = pre-mating day; Pmw = pre-mating week; Pnd = post-natal day; Pnw = post-natal week; PPS = preputial separation; Resp = respiratory; T3 = triiodothyronine; T4 = thyroxine; TSH = thyroid stimulating hormone; x = time(s); wk = week(s) D H R Emg/kg/day 10000 1r 2r 3r 12r 1000 35m 35m 10r 11r 12r 18r100 35m 39m15r36m 28r16r 8r 10r 11r 14r37m 26r 27r 28r 32r 37m 20r 34m 36m 37m 38m 17r 19r 20r 21r 10 39m 32r 33r8r 9r 13r 14r 19r 26r 27r 33r 13r 4r 5r 5r1 4r 5r 5r 0.1 24r 0.01 0.001 0.0001 c-Cat d-Dogr-Rat p-Pigq-Cow k-Monkeym-Mouse h-Rabbit a-Sheep f-Ferret j-Pigeone-Gerbil s-Hamster g-Guinea Pig n-Mink o-Other Cancer Effect Level-Animals LOAEL, More Serious-AnimalsLOAEL, Less Serious-AnimalsNOAEL - Animals Cancer Effect Level-Humans LOAEL, More Serious-HumansLOAEL, Less Serious-HumansNOAEL - Humans LD50/LC50Minimal Risk Level for effects other than Cancer eath epatic enal ndocrine 3. HEALTH EFFECTSFigure 3-2. Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - Oral Acute (≤14 days) Systemic PBDEs103BodyMetaImm NeurRepr10000 mg/kg/day 22r 12r 1000 35m 23r 42m 32r 33r 51m29r8r 9r 10r 11r 12r 18r100 39m 6r36m 43m26r 27r 28r 48r31r10r 11r 13r 14r37m 20r 41m 43m 40r 48r 49r25r 31r 38m 19r 41m 44m21r 21r 10 6r 13r 14r 7r 47r 45r1 30r 50r46r 0.1 0.01 0.001 0.0001 c-Cat d-Dogr-Rat p-Pigq-Cow k-Monkeym-Mouse h-Rabbit a-Sheep f-Ferret j-Pigeone-Gerbil s-Hamster g-Guinea Pig n-Mink o-Other Cancer Effect Level-Animals LOAEL, More Serious-AnimalsLOAEL, Less Serious-AnimalsNOAEL - Animals Cancer Effect Level-Humans LOAEL, More Serious-HumansLOAEL, Less Serious-HumansNOAEL - Humans LD50/LC50Minimal Risk Level for effects other than Cancer Weight bolic uno/Lymphor ological oductive 3. HEALTH EFFECTSFigure 3-2. Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - Oral (Continued) Acute (≥14 days) Systemic PBDEs104Devmg/kg/day 10000 1000 52r 56r 57r100 66r 78m 55r53r 66r89m 90m85m 86m 87m 88m 91h75m 79m 84m71m10 68m 53r 62r 67r72m 73m 74m 81m 65r69m 91h 62r54r 1 77m 54r 58r 59r 67r70m 75m 76m 80m 83m 65r71m 82m80m 83m 84m 61r 64r 0.1 60r 61r 63r 0.01 0.001 0.0001 c-Cat d-Dogr-Rat p-Pigq-Cow k-Monkeym-Mouse h-Rabbit a-Sheep f-Ferret j-Pigeone-Gerbil s-Hamster g-Guinea Pig n-Mink o-Other Cancer Effect Level-Animals LOAEL, More Serious-AnimalsLOAEL, Less Serious-AnimalsNOAEL - Animals Cancer Effect Level-Humans LOAEL, More Serious-HumansLOAEL, Less Serious-HumansNOAEL - Humans LD50/LC50Minimal Risk Level for effects other than Cancer elopmental 3. HEALTH EFFECTSFigure 3-2. Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - Oral (Continued) Acute (≤14 days) PBDEs105RespiCardiGastrHemaMuscHepatRenalmg/kg/day 1000 116r 116r 116r 116r 116r 118r 118r 118r 118r123r 123r 123r 123r 138m 139m 100 125r 125r 125r 125r 125r 114r 115r116r 109r 116r 112r 118r 95r111r106r 110r104r10 114r 115r97r 98r 101r 95r 116r 94r 112r 118r105r98r 101r 125r103r 137m 137m1 132m 131m 95r 0.1 123r 107r 0.01 0.001 0.0001 1E-5 1E-6 c-Cat d-Dogr-Rat p-Pigq-Cow k-Monkeym-Mouse h-Rabbit a-Sheep f-Ferret j-Pigeone-Gerbil s-Hamster g-Guinea Pig n-Mink o-Other Cancer Effect Level-Animals LOAEL, More Serious-AnimalsLOAEL, Less Serious-AnimalsNOAEL - Animals Cancer Effect Level-Humans LOAEL, More Serious-HumansLOAEL, Less Serious-HumansNOAEL - Humans LD50/LC50Minimal Risk Level for effects other than Cancer ratory ovascular ointestinal tological uloskeletal ic 3. HEALTH EFFECTSFigure 3-2. Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - Oral (Continued) Intermediate (15-364 days) Systemic PBDEs106Renal EndocriDermal Ocular Body Wmg/kg/day 1000 116r 116r 133m 134m 133m 134m 123r 123r 138m 139m 100 125r 125r 125r114r 115r 122r 93r 94r109r 116r 93r 94r 96r 120r 128r 136m 96r95r111r110r 129m108r 113r 117r10 92r 125r 128r 97r116r95r 124r93r 94r 96r 120r 121r 111r 125r117r 137m123r1 92r 130m 132m 135m 140n95r131m 131m 141n 0.1 140n142n 0.01 0.001 0.0001 1E-5 1E-6 c-Cat d-Dogr-Rat p-Pigq-Cow k-Monkeym-Mouse h-Rabbit a-Sheep f-Ferret j-Pigeone-Gerbil s-Hamster g-Guinea Pig n-Mink o-Other Cancer Effect Level-Animals LOAEL, More Serious-AnimalsLOAEL, Less Serious-AnimalsNOAEL - Animals Cancer Effect Level-Humans LOAEL, More Serious-HumansLOAEL, Less Serious-HumansNOAEL - Humans LD50/LC50Minimal Risk Level for effects other than Cancer ne eight 3. HEALTH EFFECTSFigure 3-2. Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - Oral (Continued) Intermediate (15-364 days) Systemic PBDEs107Body WMetaboImmunNeurolo100 1000 mg/kg/day 113r 113r119r10 125r123r98r 101r 102r 98r 111r 1 126r 132m 151n150m 157n127r 0.1 156r151n 107r 152r 0.01 0.001 0.0001 1E-5 1E-6 147r116r 118r 118r 148r 154r99r 100r 155r 122r 125r 149r114r 115r 144r 145r 146r116r 123r120r 121r109r 109r99r 100r 101r 102r 117r 128r 153r118r 143r108r 111r 111r110r 153r c-Cat d-Dogr-Rat p-Pigq-Cow k-Monkeym-Mouse h-Rabbit a-Sheep f-Ferret j-Pigeone-Gerbil s-Hamster g-Guinea Pig n-Mink o-Other Cancer Effect Level-Animals LOAEL, More Serious-AnimalsLOAEL, Less Serious-AnimalsNOAEL - Animals Cancer Effect Level-Humans LOAEL, More Serious-HumansLOAEL, Less Serious-HumansNOAEL - Humans LD50/LC50Minimal Risk Level for effects other than Cancer eight lic o/Lymphor gical 3. HEALTH EFFECTSFigure 3-2. Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - Oral (Continued) Intermediate (15-364 days) Systemic PBDEs108ReproductiDevelopmemg/kg/day 1000 167r167r182m 183m 208m 209m 168r 173r 100 174r165r 166r 158r 159r 170r 172r181m 158r 161r 170r 171r 177r160r 163r178m 205m 195r194r169r 198r196r 201r10 189r 204r173r 199r191r 202r171r 192r 197r162r 190r 193r 196r 200r 201r 1 179m 184m 189r 190r 204r197r 191r180m 188n 192r187n 176r 206m 203r 0.1 186m 188n 214n187n 212n 213n185m 186m 164r 207m 210m 211m 0.01 185m 0.001 164r 175r 0.0001 1E-5 1E-6 c-Cat d-Dogr-Rat p-Pigq-Cow k-Monkeym-Mouse h-Rabbit a-Sheep f-Ferret j-Pigeone-Gerbil s-Hamster g-Guinea Pig n-Mink o-Other Cancer Effect Level-Animals LOAEL, More Serious-AnimalsLOAEL, Less Serious-AnimalsNOAEL - Animals Cancer Effect Level-Humans LOAEL, More Serious-HumansLOAEL, Less Serious-HumansNOAEL - Humans LD50/LC50Minimal Risk Level for effects other than Cancer ve ntal 3. HEALTH EFFECTSFigure 3-2. Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - Oral (Continued) Intermediate (15-364 days) PBDEs109214810001000214910001000204195.9208550002002160002083100300216875015002005190003. HEALTH EFFECTSTable 3-3 Levels of Significant Exposure to Decabromodiphenyl Ether - Oral a Key to Figure Species (Strain) Exposure/ Duration/ Frequency (Route) ACUTE EXPOSURE Systemic 1 Rat (Wistar) 7 d 1 x/d System Hepatic Bd Wt NOAEL (mg/kg/day) 1000 F 1000 F Less Serious (mg/kg/day) LOAEL Serious (mg/kg/day) Reference Chemical Form Bruchajzer et al. 2010 DecaBDE (BDE209) Comments 2 Rat (Wistar) 14 d 1 x/d Hepatic Bd Wt 1000 F 1000 F Bruchajzer et al. 2010 DecaBDE (BDE209) 3 Rat (Sprague-Dawley) 14 d 1 x/d (GO) Hepatic 95.9 M Carlson 1980b DecaBDE (technical) 4 Rat Spartan once (GO) Bd Wt 5000 IRDC 1974 DecaBDE (technical) 5 Rat (Fischer- 344) 14 d 1 x/d (F) Bd Wt 16000 NTP 1986 DecaBDE (technical, 94-97% pure) 6 Rat (Long-Evans) 4 d 1 x/d (GO) Endocr 100 F Zhou et al. 2001 DecaBDE (DE-83R) No exposure-related changes in serum thyroid hormone levels. Bd Wt 300 F 7 Mouse (C57) 3 d Gd 7-9 (G) Endocr 750 F 1500 F (reduced maternal serum T4) Chi et al. 2011 DecaBDE (BDE209) 8 Mouse (B6C3F1) 14 d 1 x/d (F) Bd Wt 19000 NTP 1986 DecaBDE (technical, 94-97% pure) PBDEs110226398002368103022526.723671.345.76216715075022071.342.223. HEALTH EFFECTS9 Table 3-3 Levels of Significant Exposure to Decabromodiphenyl Ether - Oral (continued) Exposure/ Duration/a Key to Species Frequency Figure (Strain) (Route) System Bd Wt NOAEL (mg/kg/day) 9800 M 10 M b 1.34 150 b 1.34 M PBDEs111LOAEL Less Serious Serious (mg/kg/day) (mg/kg/day) 30 M (impaired learning in Pnd 25 offspring) 6.7 M (decreased spontaneous activity at 2 months) 5.76 (decreased activity and impaired habituation at 2-4 months; impaired learning at 5-7 months) 750 (10% decrease in fetal weight, 3% increase in post-implantation loss) 2.22 M (decreased activity and impaired habituation at 2 and 4 months) Reference Chemical Form Sakamoto et al. 2013 DecaBDE (BDE209) Chen et al. 2014 DecaBDE (BDE209) Viberg et al. 2007 DecaBDE (BDE209) Buratovic et al. 2014 DecaBDE Chi et al. 2011 DecaBDE (BDE209) Johansson et al. 2008 DecaBDE (BDE209) Comments No change in liver weights were observed. Mouse (C3H/HeNCrlC Developmental 10 11 12 13 14 Rat (Sprague-Dawley) Rat (Sprague-Dawley) Mouse (NMRI) Mouse (C57) Mouse (NMRI) 1 wk (F) 2 wk Gd 1-14 (GO) once Pnd 3 (G) once Pnd 3 (GO) 3 d Gd 7-9 (G) once Pnd 3 (G) 2240620223962023162.22231720.1231820.1212910003. HEALTH EFFECTSTable 3-3 Levels of Significant Exposure to Decabromodiphenyl Ether - Oral (continued) Exposure/ Duration/a Key to Species Frequency Figure (Strain) (Route) PBDEs11215 Mouse 14 d (C57BL/6J) Pnd 2-15 (IN) 16 Mouse 14 d Pnd 2-15(C57BL/6J) (IN) 17 Mouse once Pnd 3(NMRI) (G) 18 Mouse once Pnd 10(NMRI) (G) 19 Mouse once Pnd 19(NMRI) (G) INTERMEDIATE EXPOSURE Systemic 20 Rat 36 d Gd 6 - Pnd 21(Sprague-Dawley) (GO) LOAEL NOAEL Less Serious Serious System (mg/kg/day) (mg/kg/day) (mg/kg/day) 6 20 (delayed ontogeny of reflexes in males and females, increased locomotion in Pnd 70 males, reduced serum T4 in Pnd 21 males) 6 20 (learning impairment and impulsivity at 16 months of age) 2.22 M (decreased spontaneous activity at 2 and 6 months) 20.1 M 20.1 M Bd Wt 1000 F Reference Chemical Form Rice et al. 2007 DecaBDE (BDE209) Rice et al. 2009 DecaBDE (BDE209) Viberg et al. 2003b DecaBDE (BDE209) Viberg et al. 2003b DecaBDE (BDE209) Viberg et al. 2003b DecaBDE (BDE209) Biesemeier et al. 2011 DecaBDE (BDE209) Comments Compound administered via micropipette. Compound administered via micropipette. No change in spontaneous activity or habituation at 2, 4, or 6 months. No change in spontaneous activity or habituation at 2, 4, or 6 months. 215010001000215110001000219514620579090909022171003001006003. HEALTH EFFECTSTable 3-3 Levels of Significant Exposure to Decabromodiphenyl Ether - Oral (continued) PBDEs113a Key to Species Figure (Strain) 21 Rat (Wistar) 22 Rat (Wistar) 23 Rat (Sprague-Dawley) 24 Rat (CD) 25 Rat (Sprague-Dawley) Exposure/ Duration/ Frequency (Route) 21 d 1 x/d 28 d 1 x/d 31 d Gd 10 - Pnd 21 (F) 28 d (F) 33 d (Pnd 10-42) 1 x/d (G) System Hepatic Bd Wt Hepatic Bd Wt Bd Wt Hepatic Renal Endocr Bd Wt Hepatic Endocr Bd Wt LOAEL NOAEL (mg/kg/day) Less Serious (mg/kg/day) Serious (mg/kg/day) Reference Chemical Form Comments 1000 F 1000 F Bruchajzer et al. 2010 DecaBDE (BDE209) 1000 F 1000 F Bruchajzer et al. 2010 DecaBDE (BDE209) 146 F Fujimoto et al. 2011 DecaBDE (BDE209) 90 90 90 90 IRDC 1976 DecaBDE (technical) 100 M 300 M (fatty degeneration, inflammatory foci) Lee et al. 2010 DecaBDE (BDE209) 600 M 100 M (reduced serum T3) 222630030020088000800080008000800080008000800080003. HEALTH EFFECTSTable 3-3 Levels of Significant Exposure to Decabromodiphenyl Ether - Oral (continued) Exposure/ Duration/a Key to Species Frequency Figure (Strain) (Route) PBDEs11426 Rat ~11wk Pmd 21 -(Sprague-Pnd 21Dawley) (GO) 27 Rat 13 wk (Fischer- 344) (F) System Hepatic Bd Wt Resp Cardio Gastro Hemato Musc/skel Hepatic Renal Endocr Bd Wt LOAEL NOAEL (mg/kg/day) Less Serious (mg/kg/day) Serious (mg/kg/day) Reference Chemical Form Comments 300 F (hepatocyte degeneration, eosinophilic changes) Liu et al. 2012 DecaBDE (BDE209) 300 F (12% decrease in maternal body weight) 8000 8000 NTP 1986 DecaBDE (technical, 94-97% pure) 8000 8000 8000 8000 8000 8000 8000 231160606030 606022531001001001002277505022830.05203. HEALTH EFFECTSTable 3-3 Levels of Significant Exposure to Decabromodiphenyl Ether - Oral (continued) a Key to Figure Species (Strain) Exposure/ Duration/ Frequency (Route) System NOAEL (mg/kg/day) Less Serious (mg/kg/day) LOAEL Serious (mg/kg/day) Reference Chemical Form Comments 28 Rat (Wistar) 28 d (G) Hemato Musc/skel Renal Endocr Bd Wt 60 60 60 30 F 60 60 F (increased serum T3) Van der Ven et al. 2008a DecaBDE (BDE209) 29 Rat (Sprague-Dawley) 90 d 1 x/d (GO) Hepatic 100 M Wang et al. 2010 DecaBDE (BDE209) No change in kidney weight, clinical chemistry parameters, or serum thyroid hormones. Renal Endocr Bd Wt 100 M 100 M 100 M 30 Rat (Sprague-Dawley) 90 d 1 x/d (GO) Endocr 50 M Wang et al. 2011b DecaBDE (BDE209) No exposure-related changes in serum thyroid hormone levels. Bd Wt 50 M 31 Rat (Sprague-Dawley) 8 wk 7 d/wk (GO) Endocr c 0.05 M (12% increase in serum glucose) Zhang et al. 2013a DecaBDE (BDE209) Bd Wt 20 M PBDEs11522022022191602221160222316020139500950095009500950095009500950095003. HEALTH EFFECTSTable 3-3 Levels of Significant Exposure to Decabromodiphenyl Ether - Oral (continued) PBDEs116a Key to Species Figure (Strain) 32 Mouse (Tg2576) 15 d 1 x/d (GO) 33 Mouse (CD-1) 15 d 1 x/d (G) 34 Mouse (CD-1) 30 d 1 x/d (G) 35 Mouse (CD-1) 60 d 1 x/d (G) 36 Mouse 13 wk (B6C3F1) (F) Exposure/ Duration/ Frequency (Route) System Bd Wt Bd Wt Bd Wt Bd Wt Resp Cardio Gastro Hemato Musc/skel Hepatic Renal Endocr Bd Wt LOAEL NOAEL (mg/kg/day) Less Serious (mg/kg/day) Serious (mg/kg/day) Reference Chemical Form Comments 20 M Heredia et al. 2012 DecaBDE (BDE209) 160 Liang et al. 2010 DecaBDE (BDE209) 160 Liang et al. 2010 DecaBDE (BDE209) 160 Liang et al. 2010 DecaBDE (BDE209) 9500 9500 9500 9500 9500 9500 9500 9500 9500 NTP 1986 DecaBDE (technical, 94-97% pure) 226494009400237875095095022731500230815002286260232018003. HEALTH EFFECTSTable 3-3 Levels of Significant Exposure to Decabromodiphenyl Ether - Oral (continued) a Key to Figure Species (Strain) Exposure/ Duration/ Frequency (Route) System NOAEL (mg/kg/day) Less Serious (mg/kg/day) LOAEL Serious (mg/kg/day) Reference Chemical Form Comments 37 Mouse (C3H/HeNCrlC 4 wk (F) Hepatic 9400 M (moderate hepatocellular hypertrophy) Sakamoto et al. 2013 DecaBDE (BDE209) Bd Wt 9400 M 38 Mouse (Parkes) 35 d (GO) Endocr 750 M 950 M (reduced serum T3 and T4) Sarkar et al. 2015 DecaBDE (BDE209) Bd Wt 950 M 39 Mouse (CD-1) 50 d 1 x/d (GO) Bd Wt 1500 M Tseng et al. 2006 DecaBDE (BDE209) 40 Mouse (CD-1) 18 d Gd 0-17 (GO) Bd Wt 1500 F Tseng et al. 2008 DecaBDE (BDE209) 41 Mouse (BALB/c) 31 d Gd 10 - Pnd 21 (F) Bd Wt 260 F Watanabe et al. 2008 DecaBDE (BDE209) 42 Mouse (BALB/c) 28 d (F) Bd Wt 1800 F Watanabe et al. 2010a DecaBDE (BDE209) PBDEs11723499022243002009800023096023728003. HEALTH EFFECTSTable 3-3 Levels of Significant Exposure to Decabromodiphenyl Ether - Oral (continued) a Key to Figure Species (Strain) Exposure/ Duration/ Frequency (Route) System NOAEL (mg/kg/day) Less Serious (mg/kg/day) LOAEL Serious (mg/kg/day) Reference Chemical Form Immuno/ Lymphoret 43 Rat (CD) 28 d (F) 90 IRDC 1976 DecaBDE (technical) 44 Rat (Sprague-Dawley) ~11wk Pmd 21 -Pnd 21 (NS) 300 F (increased spleen weight, lesions in spleen and thymus, altered T-cell distribution, decreased serum IgM, IgG, decreased lymphocyte proliferation) Liu et al. 2012 DecaBDE (BDE209) 45 46 Rat (Fischer- 344) 13 wk (F) Rat (Wistar) 28 d (G) 8000 60 M NTP 1986 DecaBDE (technical, 94-97% pure) Van der Ven et al. 2008a DecaBDE (BDE209) 47 Mouse (C57BL/6) 10 mo every other day (GO) 800 F (impaired CD4 T cell immune function) Feng et al. 2016b DecaBDE (BDE209) PBDEs118Comments No exposure-related changes in spleen weight or spleen or thymus histology. No exposure-related changes in spleen or thymus weight or histology, no change in spleen cell subpopulations. 2014950023191800227650232920.122012022181603. HEALTH EFFECTSTable 3-3 Levels of Significant Exposure to Decabromodiphenyl Ether - Oral (continued) PBDEs119a Key to Species Figure (Strain) 48 Mouse (B6C3F1) 49 Mouse (BALB/c) Neurological 50 Rat (Sprague-Dawley) 51 Rat (Wistar) 52 Mouse (Tg2576) 53 Mouse (CD-1) Exposure/ Duration/ Frequency (Route) System NOAEL (mg/kg/day) Less Serious (mg/kg/day) LOAEL Serious (mg/kg/day) Reference Chemical Form 13 wk (F) 28 d (F) 9500 1800 F NTP 1986 DecaBDE (technical, 94-97% pure) Watanabe et al. 2010a DecaBDE (BDE209) 90 d 1 x/d (GO) 50 M Wang et al. 2011b DecaBDE (BDE209) 20 d Pnd 22-41 (G) 20.1 (decreased synaptic potency, short-term plasticity, and long-term potentiation on Pnd 60) Xing et al. 2009 DecaBDE (BDE209) 15 d 1 x/d (GO) 20 M (decreased anxiety behaviors) Heredia et al. 2012 DecaBDE (BDE209) 15 d 1 x/d (G) 160 Liang et al. 2010 DecaBDE (BDE209) Comments No change in pulmonary viral load after RSV infection 1 week after exposure period. No changes in open-field behavior. No exposure-related changes in brain weight or AchE activity. 222016022228016021281000219414623539022166003. HEALTH EFFECTSTable 3-3 Levels of Significant Exposure to Decabromodiphenyl Ether - Oral (continued) PBDEs120a Key to Species Figure (Strain) 54 Mouse (CD-1) 55 Mouse (CD-1) Reproductive 56 Rat (Sprague-Dawley) 57 Rat (Sprague-Dawley) 58 Rat (CD) 59 Rat (Sprague-Dawley) Exposure/ Duration/ Frequency (Route) System NOAEL (mg/kg/day) Less Serious (mg/kg/day) LOAEL Serious (mg/kg/day) Reference Chemical Form 30 d 1 x/d (G) 160 Liang et al. 2010 DecaBDE (BDE209) 60 d 1 x/d (G) 80 160 (transient reduction in brain AchE activity) Liang et al. 2010 DecaBDE (BDE209) 36 d Gd 6 - Pnd 21 (GO) 1000 F Biesemeier et al. 2011 DecaBDE (BDE209) 32 d Gd 10 - Pnd 21 (F) 28 d (F) 146 F 90 Fujimoto et al. 2011 DecaBDE (BDE209) IRDC 1976 DecaBDE (technical) Pnd 10-42 1 x/d (G) 600 M Lee et al. 2010 DecaBDE (BDE209) Comments No exposure-related changes in brain weight or AchE activity. No treatment-related changes in maternal toxicity, gestation length, or number of implantations. No change in number of implantations, live pups, or sex ratio. No exposure-related changes in reproductive organ weight or histology. No exposure-related changes in reproductive organ weight or testicular histology. 22253002011800020169500237775095022721005003. HEALTH EFFECTSTable 3-3 Levels of Significant Exposure to Decabromodiphenyl Ether - Oral (continued) a Key to Figure Species (Strain) Exposure/ Duration/ Frequency (Route) System NOAEL (mg/kg/day) Less Serious (mg/kg/day) LOAEL Serious (mg/kg/day) Reference Chemical Form Comments 60 Rat (Sprague-Dawley) ~11wk Pmd 21 -Pnd 21 (NS) 300 F (atrophic and fibrotic changes in ovary, decreased number of ovarian follicles) Liu et al. 2012 DecaBDE (BDE209) 61 Rat (Fischer- 344) 13 wk (F) 8000 NTP 1986 DecaBDE (technical, 94-97% pure) No exposure-related changes in reproductive organ histology. 62 Mouse (B6C3F1) 13 wk (F) 9500 NTP 1986 DecaBDE (technical, 94-97% pure) No exposure-related changes in reproductive organ histology. 63 Mouse (Parkes) 35 d (GO) 750 M 950 M (13-18% decrease in relative testis and epididymides weight, degenerative changes in seminiferous tubules, reduced sperm count/viability, reduced serum testosterone) Sarkar et al. 2015 DecaBDE (BDE209) 64 Mouse (CD-1) 50 d Pnd 21-70 (GO) 100 M 500 M (reduced amplitude of lateral head velocity of sperm; reduced sperm mitochondrial membrane potential) Tseng et al. 2006 DecaBDE (BDE209) PBDEs1212307150022852602322290021271000219323. HEALTH EFFECTSTable 3-3 Levels of Significant Exposure to Decabromodiphenyl Ether - Oral (continued) PBDEs122a Key to Species Figure (Strain) 65 Mouse (CD-1) 66 Mouse (BALB/c) 67 Mouse (BALB/c) Developmental 68 Rat (Sprague-Dawley) 69 Rat (Sprague-Dawley) Exposure/ Duration/ Frequency (Route) System NOAEL (mg/kg/day) Less Serious (mg/kg/day) LOAEL Serious (mg/kg/day) Reference Chemical Form 18 d Gd 0-17 (GO) 1500 F Tseng et al. 2008 DecaBDE (BDE209) 31 d Gd 10 - Pnd 21 (F) 260 F Watanabe et al. 2008 DecaBDE (BDE209) 31 d Gd 10 - Pnd 21 (F) 2900 F Watanabe et al. 2010b DecaBDE (BDE209) 36 d Gd 6 - Pnd 21 (GO) 1000 Biesemeier et al. 2011 DecaBDE (BDE209) 32 d Gd 10 - Pnd 21 (F) 2 (diffuse liver cell hypertrophy with cytoplasmic eosinophilia in male offspring and cytoplasmic eosinophilia in the cortical proximal tubules of the kidney of female offspring at Pnd 20; recovered by Pnw 11) Fujimoto et al. 2011 DecaBDE (BDE209) Comments No change in gestational length or litter size, no change in ovary weight. No change in the number of litters. No change in the number of litters. No treatment-related changes in litter size, sex ratio, pup survival and development, or neurobehavior or neuropathology of offspring. 23601000232520.1232620.1232720.1232820.13. HEALTH EFFECTSTable 3-3 Levels of Significant Exposure to Decabromodiphenyl Ether - Oral (continued) PBDEs123a Key to Species Figure (Strain) 70 Rat (Sprague-Dawley) 71 Rat (Wistar) 72 Rat (Wistar) 73 Rat (Wistar) 74 Rat (Wistar) Exposure/ Duration/ Frequency (Route) System NOAEL (mg/kg/day) Less Serious (mg/kg/day) LOAEL Serious (mg/kg/day) Reference Chemical Form 20 d Gd 0-19 (GO) 1000 F Hardy et al. 2002 97.34% deca-, 2.66% nona-and octa-BDE 21 d Gd 1-21 (G) 20.1 Xing et al. 2009 DecaBDE (BDE209) 61 d Gd 1 - Pnd 41 (G) 20.1 (decreased synaptic potency, short-term plasticity, and long-term potentiation in hippocampus in Pnd 60 offspring) Xing et al. 2009 DecaBDE (BDE209) 21 d Pnd 1-21 (G) 20.1 (altered long-term potentiation in hippocampus of Pnd 60 offspring) Xing et al. 2009 DecaBDE (BDE209) 19 d Pnd 3-21 (G) 20.1 (decreased synaptic potency, short-term plasticity, and long-term potentiation in hippocampus on Pnd 60) Xing et al. 2009 DecaBDE (BDE209) Comments No change in hippocampal eletrophysiological readings in Pnd 60 offspring. DecaBDE administered to dams Gd 1- Pnd 21 and to offspring Pnd 22-41. DecaBDE administered to dams. DecaBDE administered to neonates. 230510224710228434260232129029002265910023718003. HEALTH EFFECTSTable 3-3 Levels of Significant Exposure to Decabromodiphenyl Ether - Oral (continued) a Key to Figure Species (Strain) Exposure/ Duration/ Frequency (Route) System NOAEL (mg/kg/day) Less Serious (mg/kg/day) LOAEL Serious (mg/kg/day) Reference Chemical Form Comments 75 Mouse (CD-1) 18 d Gd 0-17 (GO) 10 M (hepatocytic swelling in male offspring at Pnd 71) Tseng et al. 2008 DecaBDE (BDE209) 76 Mouse (CD-1) 18 d Gd 0-17 (GO) 10 M (vacuolization in interstitial cells of testes and sperm damage in male offspring at Pnd 71) Tseng et al. 2013 DecaBDE (BDE209) 77 Mouse (BALB/c) 31 d Gd 10 - Pnd 21 (F) 34 260 (15% decrease in pup weight on Pnd 21; increased viral load in lung and IFN-gamma in BALF in offspring on Pnd 33) Watanabe et al. 2008 DecaBDE (BDE209) Pups were infected with RSV on Pnd 28. 78 Mouse (BALB/c) 31 d Gd 10 - Pnd 21 (F) 290 2900 (increased viral load in lung and altered cytokine expression BALF in offspring on Pnd 29-33) Watanabe et al. 2010b DecaBDE (BDE209) Pups were infected with RSV on Pnd 28. Cancer 79 Mouse (C3H/HeNCrlC 27 wk (F) 9100 M (CEL: liver neoplastic nodules; altered foci) Sakamoto et al. 2013 DecaBDE (BDE209) CHRONIC EXPOSURE Death 80 Mouse (C57BL/6) 2 yr every other day (GO) 800 F (5/10 died vs. 1/10 controls) Feng et al. 2015 DecaBDE (BDE209) PBDEs124203011111111113. HEALTH EFFECTSTable 3-3 Levels of Significant Exposure to Decabromodiphenyl Ether - Oral (continued) a Key to Figure Species (Strain) Exposure/ Duration/ Frequency (Route) System NOAEL (mg/kg/day) Less Serious (mg/kg/day) LOAEL Serious (mg/kg/day) Reference Chemical Form Comments Systemic 81 Rat (Sprague-Dawley) 2 yr (F) Resp Cardio 1 1 Kociba et al. 1975; Norris et al. 1975a 77% decaBDE, 22% nonaBDE Gastro 1 Hemato 1 Musc/skel 1 Hepatic 1 Renal 1 Endocr 1 Ocular 1 Bd Wt 1 PBDEs1252018255025501120 22402550255011202550255025503. HEALTH EFFECTSTable 3-3 Levels of Significant Exposure to Decabromodiphenyl Ether - Oral (continued) a Key to Figure Species (Strain) Exposure/ Duration/ Frequency (Route) System NOAEL (mg/kg/day) Less Serious (mg/kg/day) LOAEL Serious (mg/kg/day) Reference Chemical Form Comments 82 Rat (Fischer- 344) 103 wk (F) Resp Cardio 2550 F 2550 F NTP 1986 DecaBDE (technical, 94-97% pure) Gastro Hemato 1120 M 2550 F 2240 M (acanthosis) Musc/skel Hepatic 2550 F 1120 M (precancerous neoplastic nodules) Renal 2550 F Endocr 2550 F Bd Wt 2550 F PBDEs1262024778077803760 77807780778032007780320077802031120191200202577803. HEALTH EFFECTSTable 3-3 Levels of Significant Exposure to Decabromodiphenyl Ether - Oral (continued) a Key to Figure Species (Strain) Exposure/ Duration/ Frequency (Route) System NOAEL (mg/kg/day) Less Serious (mg/kg/day) LOAEL Serious (mg/kg/day) Reference Chemical Form Comments 83 Mouse (B6C3F1) 103 wk (F) Resp Cardio 7780 F 7780 F NTP 1986 DecaBDE (technical, 94-97% pure) Gastro 3760 F 7780 F (ulcers) Hemato 7780 F Musc/skel 7780 F Hepatic 3200 M (centrilobular hypertrophy and granulomas) Renal 7780 F Endocr 3200 M (follicular cell hyperplasia) Immuno/ Lymphoret 84 Rat (Sprague-Dawley) 2 yr (F) Bd Wt 7780 F 1 Kociba et al. 1975; Norris et al. 1975a 77% decaBDE, 22% nonaBDE No exposure-related changes in immune tissue histology. 85 Rat (Fischer- 344) 103 wk (F) 1200 F (splenic hematopoiesis) NTP 1986 DecaBDE (technical, 94-97% pure) 86 Mouse (B6C3F1) 103 wk (F) 7780 F NTP 1986 DecaBDE (technical, 94-97% pure) No exposure-related changes in immune tissue histology PBDEs127203311202122402550202766507780202211203. HEALTH EFFECTSTable 3-3 Levels of Significant Exposure to Decabromodiphenyl Ether - Oral (continued) a Key to Figure Species (Strain) Exposure/ Duration/ Frequency (Route) System NOAEL (mg/kg/day) Less Serious (mg/kg/day) LOAEL Serious (mg/kg/day) Comments Reference Chemical Form Reproductive 87 Rat (Sprague-Dawley) 2 yr (F) 1 M 1 F Kociba et al. 1975; Norris et al. 1975a 77% decaBDE, 22% nonaBDE 88 Rat (Fischer- 344) 103 wk (F) 2240 M 2550 F No exposure-related changes in reproductive organ histology. NTP 1986 DecaBDE (technical, 94-97% pure) 89 Mouse (B6C3F1) 103 wk (F) 6650 M 7780 F No exposure-related changes in reproductive organ histology. NTP 1986 DecaBDE (technical, 94-97% pure) Cancer 90 Rat (Fischer- 344) 103 wk (F) 1120 M (CEL: liver neoplastic nodules) NTP 1986 DecaBDE (technical, 94-97% pure) PBDEs128Table 3-3 Levels of Significant Exposure to Decabromodiphenyl Ether - Oral (continued) Exposure/ LOAEL a Key to Figure Species (Strain) Duration/Frequency (Route)System NOAEL (mg/kg/day) Reference Chemical Form Comments Less Serious (mg/kg/day) Serious (mg/kg/day) 91 Mouse (B6C3F1) 103 wk (F) 3200 M (CEL: hepatocellular adenomas and NTP 1986 DecaBDE (technical, 94-97%carcinomas) pure) 202832003. HEALTH EFFECTSPBDEs129a The number corresponds to entries in Figure 3-3. b Used to derive an acute oral minimal risk level (MRL); concentration divided by an uncertainty factor of 100 (10 for animal to human, and 10 for human variability), resulting in an MRL of 0.01 mg/kg/day. c Used to derive an intermediate oral minimal risk level (MRL); concentration divided by an uncertainty factor of 300 (3 for use of a minimal LOAEL,10 for animal to human, 10 for human variability), resulting in an MRL of 0.0002 mg/kg/day. Note on chemical form: Mixtures are identified by composition or trade name (if reported); otherwise, they are reported as "technical". Trade names for decaBDE include DE-83R (98% decaBDE). The individual congener is identified by by IUPAC number: BDE 209 = 2,2',3,3',4,4',5,5',6,6'-decaBDE. Where study authors report use of decaBDE, without further compositional information, it is assumed the pure congener (BDE 209) was used. AchE = acetylcholinesterase; BALF = brochoalveolar lavage fluid; Bd Wt = body weight; Cardio = cardiovascular; CEL = cancer effect level; d = day(s); Endocr = endocrine; (F) = feed; F = Female; (G) = gavage; Gastro = gastrointestinal; Gd = gestational day; (GO) = gavage in oil; Hemato = hematological; IFN = interferon; IgG = immunoglobulin G; IgM = immunoglobulin M; Immuno/Lymphoret = immunological/lymphoreticular; IN = ingestion ; LOAEL = lowest-observed-adverse-effect level; M = male; Musc/skel = musculoskeletal; NOAEL = no-observed-adverse-effect level; NS = not specified; Pmd = pre-mating day; Pnd = post-natal day; Pnw = post-natal week; Resp = respiratory; RSV = respiratory syncytial virus; T3 = triiodothyronine; T4 = thyroxine; x = time(s); wk = week(s); yr = year(s) Hepatic Endocrine Body Weight Developmental 100000 Systemic mg/kg/day 8m 5r 10000 9m 4r 7m 1000 1r 2r 1r 2r 7m 13m 6r 13m 100 6r3r 10r 15m 16m 18m 19m 10 10r 11r15m 16m12m 14m 17m 12m 14m 1 0.1 0.01 c-Cat d-Dogr-Rat p-Pigq-Cow k-Monkeym-Mouse h-Rabbit a-Sheep f-Ferret j-Pigeone-Gerbil s-Hamster g-Guinea Pig n-Mink o-Other Cancer Effect Level-Animals LOAEL, More Serious-AnimalsLOAEL, Less Serious-AnimalsNOAEL - Animals Cancer Effect Level-Humans LOAEL, More Serious-HumansLOAEL, Less Serious-HumansNOAEL - Humans LD50/LC50Minimal Risk Level for effects other than Cancer 3. HEALTH EFFECTSFigure 3-3. Levels of Significant Exposure to Decabromodiphenyl Ether - Oral Acute (≤14 days) PBDEs130Respiratory Cardiovascular Gastrointestinal Hematological Musculoskeletal Hepatic Renal Systemic mg/kg/day 10000 36m 36m 36m 36m 36m 36m 36m37m27r 27r 27r 27r 27r 27r 27r 1000 21r 22r 25r 26r 100 25r 29r 29r24r 24r 28r 28r 28r 10 1 0.1 0.01 0.001 0.0001 c-Cat d-Dogr-Rat p-Pigq-Cow k-Monkeym-Mouse h-Rabbit a-Sheep f-Ferret j-Pigeone-Gerbil s-Hamster g-Guinea Pig n-Mink o-Other Cancer Effect Level-Animals LOAEL, More Serious-AnimalsLOAEL, Less Serious-AnimalsNOAEL - Animals Cancer Effect Level-Humans LOAEL, More Serious-HumansLOAEL, Less Serious-HumansNOAEL - Humans LD50/LC50Minimal Risk Level for effects other than Cancer 3. HEALTH EFFECTSFigure 3-3. Levels of Significant Exposure to Decabromodiphenyl Ether - Oral (Continued) Intermediate (15-364 days) PBDEs131Endocrine Body Weight Immuno/LymphNeurological Reproductive mg/kg/day 10000 36m 36m 48m 62m37m27r 27r 45r 61r 67m 42m 49m39m 40m 65m 1000 20r 21r 22r 56r38m 38m 63m47m38m 63m25r 59r64m 26r 44r 60r41m 66m 33m 34m 35m 53m 54m 55m23r 57r 100 25r 29r 29r 64m24r 24r 43r 58r55m28r 28r 46r30r 30r 50r 28r 51r32m 31r 52m 10 1 0.1 31r 0.01 0.001 0.0001 c-Cat d-Dogr-Rat p-Pigq-Cow k-Monkeym-Mouse h-Rabbit a-Sheep f-Ferret j-Pigeone-Gerbil s-Hamster g-Guinea Pig n-Mink o-Other Cancer Effect Level-Animals LOAEL, More Serious-AnimalsLOAEL, Less Serious-AnimalsNOAEL - Animals Cancer Effect Level-Humans LOAEL, More Serious-HumansLOAEL, Less Serious-HumansNOAEL - Humans LD50/LC50Minimal Risk Level for effects other than Cancer or 3. HEALTH EFFECTSFigure 3-3. Levels of Significant Exposure to Decabromodiphenyl Ether - Oral (Continued) Intermediate (15-364 days) Systemic PBDEs132Developmental Cancer * mg/kg/day 10000 1000 100 10 1 0.1 0.01 0.001 0.0001 79m 78m 68r 70r 78m 71r 72r 73r 74r 69r *Doses represent the lowest dose tested per study that produced a tumorigenicresponse and do not imply the existence of a threshold for the cancer endpoint.c-Cat d-Dogr-Rat p-Pigq-Cow k-Monkeym-Mouse h-Rabbit a-Sheep f-Ferret j-Pigeone-Gerbil s-Hamster g-Guinea Pig n-Mink o-Other Cancer Effect Level-Animals LOAEL, More Serious-AnimalsLOAEL, Less Serious-AnimalsNOAEL - Animals Cancer Effect Level-Humans LOAEL, More Serious-HumansLOAEL, Less Serious-HumansNOAEL - Humans LD50/LC50Minimal Risk Level for effects other than Cancer 77m 77m 75m 76m 3. HEALTH EFFECTSFigure 3-3. Levels of Significant Exposure to Decabromodiphenyl Ether - Oral (Continued) Intermediate (15-364 days) PBDEs133Death RespiratoCardiovaGastrointHematoloMusculosHepatic Renal EndocrinOcular Body WeiImmuno/ReproducCancer * 10000 mg/kg/day 83m 83m 83m 83m 83m 83m 83m 86m 89m89m 83m 83m 83m 91m 82r 82r 82r 82r 82r 82r 82r 88r82r 88r 85r82r 82r 90r1000 80m 100 10 1 81r 81r 81r 81r 81r 81r 81r 81r 81r 81r 84r 87r 87r *Doses represent the lowest dose tested per study that produced a tumorigenicresponse and do not imply the existence of a threshold for the cancer endpoint.0.1 c-Cat d-Dogr-Rat p-Pigq-Cow k-Monkeym-Mouse h-Rabbit a-Sheep f-Ferret j-Pigeone-Gerbil s-Hamster g-Guinea Pig n-Mink o-Other Cancer Effect Level-Animals LOAEL, More Serious-AnimalsLOAEL, Less Serious-AnimalsNOAEL - Animals Cancer Effect Level-Humans LOAEL, More Serious-HumansLOAEL, Less Serious-HumansNOAEL - Humans LD50/LC50Minimal Risk Level for effects other than Cancer ry scular estinal gical keletal e ght Lymphor tive 3. HEALTH EFFECTSFigure 3-3. Levels of Significant Exposure to Decabromodiphenyl Ether - Oral (Continued) Chronic (≥365 days) Systemic PBDEs134 PBDEs 1353. HEALTH EFFECTSIn intermediate-duration dietary studies with decaBDE, there was no exposure-related mortality in rats that were exposed to estimated dietary doses of ≤90 mg/kg/day for 28 days (IRDC 1976) or rats and micefed estimated doses of ≤8,000 and ≤9,500 mg/kg/day, respectively, for 13 weeks (NTP 1986). In chronicstudies, there were no effects on survival in rats that were fed 0.01–1.0 mg/kg/day of a 77.4% decaBDEmixture (containing 21.8% nonaBDE and 0.8% octaBDE) for 2 years (Kociba et al. 1975; Norris et al.1975a), or in rats and mice fed decaBDE in estimated doses of ≤2,550 and ≤7,780 mg/kg/day,respectively, for 103 weeks (NTP 1986). However, Feng et al. (2015) reported that 5/10 female micedied following exposure to decaBDE (98% purity) every other day via gavage for up to 2 years at a doseof 800 mg/kg/dose, compared with 1/10 vehicle control mice. No cause of death or gross or microscopicpathology was reported for animals that died. In surviving mice, various histopathological lesions were qualitatively described in the brain, heart, lung, liver, spleen, kidney, and ovaries of exposed animals, butincidence data were not provided. Due to the high mortality and lack of quantitative data, this chronicstudy is not discussed in the Systemic Effects section below.The LD50 and LOAEL values for death are recorded in Tables 3-2 (lower BDEs) and 3-3 (decaBDE) and plotted in Figures 3-2 (lower BDEs) and 3-3 (decaBDE).3.2.2.2 Systemic EffectsThe systemic effects in humans and animals following oral exposure to PBDEs are described below. Thehighest NOAEL and all LOAEL values from each reliable study for systemic end points in each species and duration category are recorded in Tables 3-2 (lower-brominated PBDEs) or 3-3 (decaBDE) and plotted in Figures 3-2 (lower-brominated PBDEs) or 3-3 (decaBDE).Respiratory Effects. No studies were located regarding respiratory effects in humans after oralexposure to PBDEs. Effects of PBDEs on respiratory function have not been studied in orally exposed animals.No histopathological changes in the respiratory tract or exposure-related changes in lung weight were observed in rats exposed to pentaBDE at doses up to 250 mg/kg/day for 28 days via gavage (Oberg et al. 2010; Van der ven et al. 2008a) or in rats exposed to dietary pentaBDE at doses up to 100 mg/kg/day for90 days (WIL Research Laboratories 1984). Similarly, no changes in respiratory tract histology were observed in rats exposed to dietary octaBDE at doses up to 750 mg/kg/day for 13 weeks (IRDC 1977). PBDEs 1363. HEALTH EFFECTSNo exposure-related changes in lung weight were observed in rats exposed to a dietary penta-decaBDEmixture (52.1% pentaBDE, 44.2% decaBDE, 0.4% octaBDE) at 20 mg/kg/day for 70 days (Ernest et al. 2012) or in F0 or F1 rats exposed pentaBDE at 18 mg/kg/day from GD 6 to PND 18 via gavage (Ellis-Hutchings et al. 2006).To determine if PBDE exposure alters vitamin A homeostasis in rats in a manner similar to the relatedPBB compounds (ATSDR 2004), vitamin A levels were measured in lung tissue following exposure to pentaBDE at 0, 2.5, 25, or 250 mg/kg/day via gavage (Oberg et al. 2010). No significant changes in lung vitamin A levels were observed at any dose.No histopathological changes in respiratory tract tissues were found in rats and mice exposed to dietarydecaBDE at estimated doses of ≤8,000 and ≤9,500 mg/kg/day, respectively, for 13 weeks or estimateddoses of ≤2,550 and ≤7,780 mg/kg/day, respectively, for 103 weeks (NTP 1986). Additionally, no histopathological changes in respiratory tract tissues were observed in rats that were fed ≤1.0 mg/kg/day of a 77.4% decaBDE mixture (containing 21.8% nonaBDE and 0.8% octaBDE) for 2 years (Kociba et al.1975; Norris et al. 1975a).Based on animal studies, respiratory effects are not likely to occur following oral exposure to PBDEs.Cardiovascular Effects. No relationship was found between serum concentrations of tetraBDE(BDE 47) and carotid atherosclerosis (carotid artery plaques determined by ultrasound) or stroke in apopulation of 1,016 70-year-old volunteers (50.2% female) from Uppsala, Sweden (Lee et al. 2012; Lind et al. 2012). BDE 47 was detected in the serum of 77.2% of subjects with a median concentration of12.6 pg/mL. No associations were observed between serum PBDE levels and gestational hypertension in258 pregnant women from the Longitudinal Investigation of Fertility and the Environment (LIFE) cohortin Michigan and Texas (Smarr et al. 2016).A case control study of 43 children (9–11 years old) from Upstate New York reported an associationbetween higher levels of serum PBDEs and greater sympathetic activation during acute psychologicalstress and greater anger (Gump et al. 2014). Specifically, BDE 28 was associated with greater heart rate,lower pre-ejection period, and lower total peripheral resistance. BDE 47 and BDE 100 were associated with significantly lower diastolic blood pressure, and BDE 100 was also associated with a shorter preejection period during acute stress. Lipid-adjusted mean blood levels of BDEs 28, 47, and 100 were 1.07, 8.53, and 0.86 ng/g, respectively. The investigators speculated that PBDE-induced increased levels of PBDEs 1373. HEALTH EFFECTScalcium/calmodulin-dependent protein kinase II in the hippocampus may mediate the heart’s response toβ-adrenergic stimulation.Effects of PBDE on cardiovascular function have not been studied in orally exposed animals. No exposure-related changes in heart histology or weight were observed in rats exposed to pentaBDE atdoses up to 250 mg/kg/day for 28 days via gavage (Oberg et al. 2010; Van der ven et al. 2008a), dietarypentaBDE at doses up to 100 mg/kg/day for 90 days (WIL Research Laboratories 1984), or dietaryoctaBDE at doses up to 750 mg/kg/day for 13 weeks (IRDC 1977). Additionally, no exposure-relatedchanges in heart weight were observed in rats exposed to dietary penta- or octaBDE at doses up to90 mg/kg/day for 28 days (IRDC 1976), in rats exposed to a dietary penta-decaBDE mixture (52.1%pentaBDE, 44.2% decaBDE, 0.4% octaBDE) at 20 mg/kg/day for 70 days (Ernest et al. 2012), in F0 orF1 rats exposed to dietary pentaBDE at doses up to 0.31 mg/kg/day for 4 weeks prior to mating throughPNW 6 or 33 (Zhang et al. 2009), in F0 or F1 rats exposed pentaBDE at 18 mg/kg/day from GD 6 toPND 18 via gavage (Ellis-Hutchings et al. 2006), or in rats exposed to diBDE at 1.2 mg/kg/day for28 days via gavage (Zhang et al. 2014).No histopathological changes in the heart were found in rats and mice exposed to dietary decaBDE atestimated doses of ≤8,000 and ≤9,500 mg/kg/day, respectively, for 13 weeks or estimated doses of≤2,550 and ≤7,780 mg/kg/day, respectively, for 103 weeks (NTP 1986). No exposure-related changes inheart histology or weight were observed in rats that were fed ≤1.0 mg/kg/day of a 77.4% decaBDEmixture (containing 21.8% nonaBDE and 0.8% octaBDE) for 2 years (Kociba et al. 1975; Norris et al. 1975a). In addition, no exposure-related changes in heart weight were observed in rats exposed decaBDEat doses up to 20 mg/kg/day for 8 weeks via gavage (Zhang et al. 2013a) or to dietary decaBDE at doses up to 90 mg/kg/day for 28 days (IRDC 1976).Based on human and animal studies, cardiovascular effects are not likely to occur following oral exposureto PBDEs.Gastrointestinal Effects. No studies were located regarding gastrointestinal effects in humans afteroral exposure to PBDEs.No histopathological changes in the gastrointestinal tract were found in rats exposed to pentaBDE atdoses up to 200 mg/kg/day for 28 days via gavage (Van der ven et al. 2008b), dietary pentaBDE at doses PBDEs 1383. HEALTH EFFECTSup to 100 mg/kg/day for 90 days (WIL Research Laboratories 1984), or dietary octaBDE at doses up to750 mg/kg/day for 13 weeks (IRDC 1977). No histopathological changes in gastrointestinal tract tissues were found in rats and mice fed decaBDE inestimated doses of ≤8,000 and ≤9,500 mg/kg/day, respectively, for 13 weeks (NTP 1986). In chronicdietary studies, there was no gastrointestinal tract histopathology in rats that were fed ≤1.0 mg/kg/day of a77.4% decaBDE mixture (containing 21.8% nonaBDE and 0.8% octaBDE) for 2 years (Kociba et al.1975; Norris et al. 1975a). Higher dietary doses of decaBDE for 103 weeks caused acanthosis of the forestomach in rats exposed to 2,240 mg/kg/day (no effects at ≤1,200 mg/kg/day) and stomach ulcers in mice exposed to 7,780 mg/kg/day (no effects at ≤3,760 mg/kg/day) (NTP 1986).Based on animal studies, gastrointestinal effects are not likely to occur following oral exposure to PBDEs at environmentally-relevant exposure concentrations.Hematological Effects.Human Studies. Hematological end points were evaluated in a subset of 18 of a cohort of 33 children (18 girls and 15 boys) born in the Amsterdam/Zaandam area of the Netherlands and aged 14–19 years atthe time of the study (Leijs et al. 2009). Serum PBDE concentrations (determined as the sum ofcongeners 28, 47, 85, 99, 100, 153, 154, and 183) ranged from 5 to 74 ng/g lipid with a mean of 13.9 ng/glipid. Serum samples were used to assess hemoglobin, thrombocytes, and white blood cell count and differential. The researchers found a negative statistical association (not further described) between number of lymphocytes and PBDE concentrations in serum. Congener-specific analysis showed the maincontributors to be BDE 183, BDE 154, and BDE 85. Although dioxins and PCBs were also assessed, no association with lymphocytes was found for these chemicals. The only other finding for PBDE was a positive association between serum hemoglobin and PBDE, primarily due to congeners 85 and 153. A Chinese study of 40 subjects exposed to PBDEs in an electronic waste dismantling area reported a significantly elevated neutrophil percentage among the workers compared to a group of 15 unexposedcontrols (Xu et al. 2015a). In addition, exposed subjects had significantly lower percentages ofmonocytes, lymphocytes, hemoglobin, and platelets than controls, while total white cell counts were notsignificantly different between the two groups. PBDEs assessed included congeners 15, 27, 85, 99, 100, 153, 154, and 184. Exposure was likely by multiple routes. PCBs also were found associated withlowered percentages of monocytes and lymphocytes. PBDEs 1393. HEALTH EFFECTSAnimal StudiesLower-brominated PBDEs: Following exposure to dietary octaBDE at 750 mg/kg/day for 13 weeks, female rats showed 26, 22, and 22% decreases in erythrocytes, hematocrit, and hemoglobin, respectively;male rats showed a 10% decrease in erythrocytes at 600 mg/kg/day (IRDC 1977). No changes wereobserved in total or differential white blood cell counts (IRDC 1977). No changes in hematologicalparameters, including erythrocyte count, hemoglobin, hematocrit, full and differential leukocyte count, and platelet count, were observed in rats following pentaBDE exposure to dietary doses up to100 mg/kg/day for 90 days (WIL Research Laboratories 1984) or gavage up to 250 mg/kg/day for 28 days(Oberg et al. 2010). In another 28-day study, rats were exposed to pentaBDE at 0, 0.27, 0.82, 2.47, 7.4, 22.2, 66.7, or200 mg/kg/day via gavage (Van der ven et al. 2008b). The study authors reported minor dose-relatedchanges in white blood cell differentials from femoral shaft bone marrow in male rats (blood hematologynot evaluated); however, the doses at which these effects were observed were not reported. Instead,results were reported in terms of benchmark dose (BMD) analysis (Van der ven et al. 2008b).Statistically significant changes included increased number and percentage of monocytes (maximum increase of 69.5 and 66.7%, respectively; BMD/BMDLRD20%=11.2/0.7 and 31.8/3 mg/kg/day, respectively) and decreased percentage of eosinophilic granulocytes (maximum decrease of 20%; (BMDRD20%=28.6, BMDL not determined). The study authors also reported a significant, dose-relatedincrease in the number and percentage of large unstained cells (maximum increase of 79.7 and 85.1%, respectively; BMD/BMDLRD20%=76.3/42.8 and 64.3/9.8 mg/kg/day, respectively), which they indicatedmay represent large or reactive lymphocytes, monocytes, or leukemic blasts (abnormal myeloblasts), typically associated with viral disease, leukemia, or endocrine active compounds. Since results were reported in terms of BMD analysis only, data and statistics for individual dose groups were not available for independent analysis. No data regarding other standard hematological end points were reported (Vander ven et al. 2008b). Male mink exposed to dietary pentaBDE at doses of 0.63 or 0.78 mg/kg/day for 9 weeks showed significant 13 and 12% decreases in hematocrit, respectively; neither the number of red blood cells norhemoglobin levels were evaluated (Martin et al. 2007). According to study authors, most of the ranges in the differential white blood cell counts fell within that expected for male mink of this age; however, thepercentage of neutrophils was increased significantly by ~22% at 0.63 mg/kg/day and 37% at0.78 mg/kg/day, the percentage of lymphocytes decreased significantly by ~33% at 0.78 mg/kg/day, and PBDEs 1403. HEALTH EFFECTSthe neutrophil:lymphocyte ratio increased significantly by ~90% at 0.78 mg/kg/day (data reported graphically) (Martin et al. 2007). Total white blood cell counts were not reported. No hematologicaleffects were observed at 0.08 mg/kg/day (Martin et al. 2007). The minor hematological changes are ofuncertain toxicological significance.In a poorly-reported study, mouse dams exposed to pentaBDE from GD 6 to PND 21 via gavage did notshow any hematological effects at doses up to 200 mg/kg/day (Hong et al. 2010). DecaBDE: No hematological changes were observed in rats exposed to ≤60 mg/kg/day of decaBDE viagavage for 28 days (Van der ven et al. 2008a). In dietary studies with decaBDE, no hematologicalchanges were found in rats exposed to ≤800 mg/kg/day for 30 days (Norris et al. 1973, 1975a), ≤8,000 mg/kg/day for 13 weeks (NTP 1986), or ≤2,550 mg/kg/day for 103 weeks (NTP 1986), or in miceexposed to ≤9,500 mg/kg/day for 13 weeks or ≤7,780 mg/kg/day for 103 weeks (NTP 1986). There alsowere no hematological effects in rats exposed by diet to ≤1.0 mg/kg/day of a 77.4% decaBDE mixture(containing 21.8% nonaBDE and 0.8% octaBDE) for 2 years (Kociba et al. 1975; Norris et al. 1975a).In a poorly-reported study, mouse dams exposed to decaBDE from GD 6 to PND 21 via gavage showed significant increases in the number of white blood cells and neutrophils at 500 mg/kg/day, compared with controls, but not at 2,500 or 12,500 mg/kg/day (Hong et al. 2010). Summary. Minor hematological changes observed in humans and animal are of uncertain toxicologicalsignificance. Based on the available information, it is unlikely that adverse effects would occur in the human hematological system following oral PBDE exposure.Musculoskeletal Effects. No association was found between serum PBDE concentrations (BDE 47, BDE 99, BDE 100, BDE 153, BDE 154, BDE 209, and their sum) and bone mineral density (measured byx-ray in the forearm) or serum concentrations of biochemical markers of bone metabolism (osteocalcinand crosslaps [a peptide of type I collagen]) in a population of 50 post-menopausal Swedish women married or previously married to professional fisherman on the east coast of the country and born before 1954 (Weiss et al. 2006). Median serum ∑PBDE concentration in this population was 3.6 ng/g fat.No musculoskeletal changes were seen in rats following exposure to pentaBDE at doses up to200 mg/kg/day for 28 days via gavage (Van der ven et al. 2008b) or dietary pentaBDE at doses up to 100 mg/kg/day for 90 days (WIL Research Laboratories 1984). PBDEs 1413. HEALTH EFFECTSNo changes in bone parameters were observed in rats exposed to decaBDE at doses up to 60 mg/kg/dayfor 28 days via gavage (Van der ven et al. 2008a). Dietary studies with decaBDE found no histopathological changes in musculoskeletal tissues in rats exposed to ≤8,000 mg/kg/day for 13 weeks (NTP 1986), ≤1.0 mg/kg/day (77.4% containing 21.8% nonaBDE and 0.8% octaBDE) for 2 years (Kociba et al. 1975; Norris et al. 1975a), or ≤2,550 mg/kg/day for 103 weeks (NTP 1986), or in miceexposed to ≤9,500 mg/kg/day for 13 weeks or ≤7,780 mg/kg/day for 103 weeks (NTP 1986). Based on human and animal studies, musculoskeletal effects are not likely to occur following oralexposure to PBDEs.Hepatic Effects. No studies were located regarding hepatic effects in humans after oral exposure toPBDEs.Acute-Duration Animal StudiesLower-brominated PBDEs: A series of gavage studies evaluated liver histology following a singleexposure to pentaBDE doses of 0, 25, 200, or 2,000 mg/kg or repeat exposures to pentaBDE doses of 0, 8, 40, or 200 mg/kg/day for 7 or 14 days (Bruchajzer et al. 2010, 2011). Fatty degeneration of the liver wasobserved in female rats following pentaBDE exposure via gavage at 2,000 mg/kg for 1 day (Bruchajzer etal. 2011) or 200 mg/kg/day for 7 or 14 days (Bruchajzer et al. 2010). In the single-exposure study, ratsfrom the 2,000 mg/kg group showed steatosis of the microvesicular type, which was most frequentlyobserved in the central and intermediate zones of lobules; however, animal incidence numbers were notreported (Bruchajzer et al. 2011). At 4 and 12 hours after pentaBDE administration, these changes werenoted in no more than 25% of hepatocytes; after 24–120 hours, fatty degeneration embraced 26–75% ofhepatocytes; and after 120 hours, a mixed type of change (micro- and macrovesicular) was observed(Bruchajzer et al. 2011). In the repeated-exposure studies, rats from the 200 mg/kg/day group showed steatosis of the microvesicular type in 3–25% of hepatocytes after 7 days and steatosis of the microvesicular and mascrovesicular type in 26–75% of hepatocytes in the central and intermediate zones of lobules (Bruchazjer et al. 2010). Again, animal incidence data were not reported. No exposure-relatedhistological changes in the liver were reported for single doses ≤200 mg/kg or repeated doses≤40 mg/kg/day (Bruchazjer et al. 2010, 2011). In the only other study that included histopathologicalexamination, histological changes were not observed in male rats 45 days after a single administration ofpentaBDE at doses up to 1.2 mg/kg (Albina et al. 2010). PBDEs 1423. HEALTH EFFECTSIn the Bruchazjer et al. (2010, 2011) studies, no biologically relevant changes were observed in serumclinical chemistry values. Serum alanine transaminase (ALT) and aspartate transaminase (AST) activitylevels were slightly increased to 130 and 170–190% of control values following a single gavage exposureto 2,000 mg/kg, respectively, after 72–120 hours; no changes were reported for doses ≤200 mg/kg(Bruchazjer et al. 2011). The study authors did not report statistics; however, these serum chemistrychanges are not considered biologically relevant since the magnitude of change, compared with control, is<2-fold. Similarly, no exposure-related changes were observed in serum ALT or AST at doses up to 200 mg/kg/day for 7 or 14 days (Bruchajzer et al. 2010). Another 14-day gavage study found no exposure-related changes in serum sorbital dehydrogenase (SDH) levels in rats exposed to pentaBDE at56.4 mg/kg/day or octaBDE at 76.6 mg/kg/day (equimolar doses) (Carlson 1980b). In male rats exposed once to pentaBDE at doses of 0, 0.6, or 1.2 mg/kg, serum ALT activity was significantly increased by~72% in the 1.2 mg/kg group, compared with control, when evaluated 45 days after exposure (Alonso etal. 2010). However, this change is not considered biologically relevant because the magnitude of change,compared with control, is <2-fold, the measured ALT activity of 31 U/l was well within reference valueranges for adult male rats (20–81 U/l) (Charles River Laboratories 1998), and no changes were observed in serum AST, alkaline phosphatase (ALP), or lactate dehydrogenase (LDH) (Alonso et al. 2010).Bruchajzer et al. (2010, 2011) also reported altered markers of oxidative stress in the liver followingexposure to pentaBDE. Liver malondialdehyde (MDA) levels were significantly elevated after exposure to ≥200 mg/kg/day for 1–14 days (Bruchajzer et al. 2010, 2011). Levels of reduced glutathione (GSH) in the liver were not increased following a single exposure; however, GSH concentration was elevated at≥40 mg/kg/day after 7 days and ≥8 mg/kg/day after 14 days (Bruchajzer et al. 2010, 2011). Oxidized glutathione (GSSG) was significantly elevated after single exposures to ≥25 mg/kg, but no changes in glutathione S-transferase (GST) activity were observed (Bruchajzer et al. 2010). Liver GSSG levels andGST activity were not evaluated following 7- or 14-day exposures. In another study, liver GSH levelswere significantly decreased and liver SOD activity, GSSG levels, and GSSG/GSH ratio weresignificantly increased in male rats 45 days after a single gavage administration of pentaBDE at 0.6 or1.2 mg/kg (Albina et al. 2010).In two other studies by Bruchazjer and colleagues (Bruchazjer 2011; Bruchazjer et al. 2012), female rats were examined for hepatic porphyria following exposure to pentaBDE at doses of 0, 8, 40, or200 mg/kg/day or octaBDE at doses of 0, 2, 8, 40, or 200 mg/kg/day via gavage for 7 or 14 days. PentaBDE caused significant dose-related elevations in total porphyrin levels in the liver following PBDEs 1433. HEALTH EFFECTSexposure to ≥40 mg/kg/day for 7 days (~2–6-fold) or ≥8 mg/kg/day for 14 days (~3–11-fold). Following gavage exposure to octaBDE for 7 days, liver concentrations of high carboxylated porphyrins (octa- and hepta-) were significantly elevated at 8, 40, and 200 mg/kg/day by ~4-, 4-, and 7-fold, respectively,compared with vehicle controls (Bruchajzer et al. 2012). Lower carboxylated porphyrins were <1% oftotal liver porphyrins measured, and were not further analyzed. Following exposure to octaBDE for14-days, total liver concentrations of high carboxylated porphyrins were increased by ~3-fold in the 200 mg/kg/day group only, compared with vehicle controls (Bruchajzer et al. 2012). However, thevehicle controls at this duration had an unusually high hepatic porphyrin levels (4-fold increase compared with untreated control). When compared with the untreated controls, porphyrin levels were significantlyincreased by ~2-, 4-, 6-, and 7-fold in the 2, 8, 40, and 200 mg/kg/day groups, respectively, supportingthat exposure to octaBDE for 14 days causes porphyria as observed in the 7-day study. These studies alsomeasured hepatic delta-aminolevulinate synthase (ALA-S) activity and delta-aminolevilinate dehydratase(ALA-D) activity, two enzymes involved in heme biosynthesis. The results were mixed. PentaBDEexposure caused a significant elevation in ALA-S activity at ≥40 mg/kg/day after 7 or 14 days (ALA-D was not measured), while octaBDE exposure caused a significant increase in ALA-S activity at≥8 mg/kg/day after 7 days and significant decreases in ALA-S activity at 200 mg/kg/day after 14 daysand ALA-D activity at ≥40 mg/kg/day after 7 or 14 days (Bruchajzer 2011; Bruchajzer et al. 2012). The gavage studies by Bruchajzer and colleagues (Bruchajzer 2011; Bruchajzer et al. 2012) also reported elevated liver weights accompanied by hepatic microsomal enzyme induction. Significant relative liverweight increases (data reported graphically) were about ≥30% after a single exposure to pentaBDE at≥200 mg/kg (Bruchajzer et al. 2011), ≥13% after exposure to pentaBDE at 8 or 200 mg/kg/day for 7 days(Bruchajzer et al. 2010), or ≥25% after exposure to pentaBDE at ≥40 mg/kg/day for 14 days (Bruchazjeret al. 2010). Hepatic enzyme induction (e.g., increased CYP1A and CYP2B activity and/or CYP1A1, CYP4A, and total cytochrome P450 protein levels) was observed at lower doses: ≥25 mg/kg in the singledose study and ≥8 mg/kg/day in the 7- and 14-day studies (Bruchazjer et al. 2010, 2011).Consistent with the findings of Bruchazjer and colleagues (Bruchajzer 2011; Bruchajzer et al. 2012), elevated liver weights have been reported following acute exposure to penta-, octa-, or tetraBDE in several other animal studies. Significantly increased liver weights were reported in rats and mice exposedto penta-, tetra-, or octaBDE via gavage at ≥10 and ≥18 mg/kg/day, respectively, for 1–14 days (Carlson 1980b; Fowles et al. 1994; Hallgren et al. 2001; Mercado-Feliciano and Bigsby 2008a; Richardson et al. 2008; Stoker et al. 2004, 2005; Zhou et al. 2001). However, no exposure-related changes in relative liverweight were observed in female rats exposed to tetraBDE at doses up to18 mg/kg/day for 14 days via PBDEs 1443. HEALTH EFFECTSgavage (Hallgren and Darnerud 2002). In the studies that evaluated hepatic enzyme induction (e.g.,elevated EROD, MROD, PROD, and UDPGT activity and/or increased CYP protein levels), significantlyincreased enzyme levels and activities were observed in rats and mice at gavage doses of ≥6 and≥3 mg/kg/day, respectively, and were always observed at doses at or below the dose causing elevatedliver weights in the same study (Carlson 1980b; Fowles et al. 1994; Hallgren and Darnerud 2002;Hallgren et al. 2001; Richardson et al. 2008; Stoker et al. 2004; Zhou et al. 2001). Exposure to a singlepentaBDE dose of 0.03 or 0.6 mg/kg on GD 6 via gavage caused significant induction of hepatic enzymes (EROD, UDPGT) in dams and pups from the 0.6 mg/kg group on PND 22; liver weight was not evaluated (Kuriyama et al. 2007). Collectively, these findings support that elevated liver weight following exposureto lower-brominated PBDEs is associated with hepatic enzyme induction.Richardson et al. (2008) evaluated genomic changes in mice following exposure to tetraBDE at 0, 3, 10, or 100 mg/kg/day for 4 days via gavage. The mRNA levels of several hepatic enzymes were significantlyelevated at ≥20 mg/kg/day (Cyp2b10, Ugt1a1, Ugt1a7, Ugt2b2). In addition, several exposure-relatedchanges were observed in hepatic efflux transporter (Mrp3, Mdr1a) and thyroid hormone transporter (Ttr,Mct8) mRNA expression levels in hepatic tissue (Richarson et al. 2008).To determine if PBDE exposure alters vitamin A homeostasis in a manner similar to the related PBBs(ATSDR 2004), hepatic vitamin A levels were measured in mice and rats exposed to pentaBDE at 0, 18,or 36 mg/kg/day and mice exposed to tetraBDE at 0 or 18 mg/kg/day for 14 days via gavage (Hallgren etal. 2001). In pentaBDE-exposed animals, hepatic vitamin A levels were significantly decreased by24 and 26% in rats in the 18 and 36 mg/kg/day groups, respectively, and 22% in mice in the36 mg/kg/day group. No changes in hepatic vitamin A levels were observed in mice exposed to penta- ortetraBDE at 18 mg/kg/day. DecaBDE: Unlike the lower-brominated PBDEs, there is no evidence of hepatic toxicity following acute exposure to decaBDE. Exposure to decaBDE at doses up to 1,000 mg/kg/day for 7 or 14 days via gavagedid not produce liver damage, as indicated by liver histology or clinical chemistry, nor did it produceincreased liver weight or hepatic enzyme induction in female rats (Bruchajzer et al. 2010). Carlson (1980b) observed no changes in serum SDH activity in rats exposed to decaBDE at 95.9 mg/kg/day for14 days via gavage; significantly elevated liver weights were found, but in the absence of hepatic enzyme induction. In a shorter-duration study, exposure to decaBDE at doses up to 100 mg/kg/day for 4 days viagavage did not cause changes in liver weight or hepatic enzyme induction in female rats; clinicalchemistry and histology were not examined (Zhou et al. 2001). PBDEs 1453. HEALTH EFFECTSSakamoto et al. (2013) evaluated both liver weight and hepatocyte proliferation in female rats followingexposure to decaBDE at 980 mg/kg/day for 1 week. Consistent with other studies, no exposure-relatedfindings in absolute or relative liver weight were observed. Additionally, no exposure-related changes inhepatocyte proliferation were observed, as evaluated by labeling indices of proliferating cell nuclearantigen (PCNA) in paraffin-embedded liver sections.Intermediate-Duration Animal StudiesLower-brominated PBDEs: Histopathological effects in the liver have been consistently observed in animals exposed to lower-brominated PBDEs for 15–90 days at doses as low as 2 mg/kg/day.Hepatocytomegaly (hepatocyte hypertrophy) was observed in rats exposed to dietary pentaBDE at doses≥2 mg/kg/day (lowest dose tested) for 13 weeks (WIL Research Laboratories 1984). Thehepatocytomegaly was dose-related with respect to severity (some affected hepatocytes at higher doses had vacuoles that likely contained lipid) and was not completely reversible, as it was still evident in≥10 mg/kg/day males and 100 mg/kg/day females at 24 weeks postexposure in lessened severity andincidence. Females exposed to 2 or 100 mg/kg/day pentaBDE for 90 days also had an increasedincidence of degeneration and necrosis of individual liver parenchymal cells at 24 weeks postexposure;the investigators concluded that this may represent the final loss of previously damaged cells andprobably should be considered compound-related (WIL Research Laboratories 1984). Similarly, dietaryexposure to octaBDE at 0, 100, 1,000, or 10,000 ppm (0, 5, 50, or 600 mg/kg/day in males and 0, 7, 70, or750 mg/kg/day in females) for 13 weeks caused liver lesions in 40% of males at 5 mg/kg/day and 100%of both sexes at ≥50–70 mg/kg/day (IRDC 1977). The lesions were dose-related in severity as well as incidence and characterized by cytomegaly, change in hepatocytic cytoplasm to a finely granular, homogeneous type, and cytoplasmic vacuolation. At 600–750 mg/kg/day, many of the livers hadvacuolation of centrolobular hepatocytes and some had hepatocyte necrosis. Examinations performed at8 weeks and 6 months postexposure showed that the liver effects persisted in the rats exposed to ≥50– 70 mg/kg/day for 13 weeks (IRDC 1977).Hepatocellular hypertrophy was also observed in rats exposed to dietary penta- or octaBDE at≥9 mg/kg/day for 28 days (IRDC 1976), in rats exposed to pentaBDE at ≥3 mg/kg/day for 15–28 days viagavage (Becker et al. 2012; Fattore et al. 2001; Oberg et al. 2010), and in rats exposed to tetraBDE at150 mg/kg/day via gavage for 12 weeks (only dose tested) (Zhang et al. 2015a, 2015b). Nohistopathological changes were observed in mice exposed to tetraBDE at 1 mg/kg/day via gavage for PBDEs 1463. HEALTH EFFECTS6 weeks (only dose tested) (McIntyre et al. 2015). In another gavage study, rats were exposed to pentaBDE at doses of 0, 0.27, 0.82, 2.47, 7.4, 22.2, 66.7, or 200 mg/kg/day for 28 days (Van der ven et al. 2008b). The study authors reported centrilobular hypertrophy and an increased ratio of binucleated hepatocytes; however, the incidence data and dose(s) at which effects were observed were not reported (Van der ven et al. 2008b). Van der ven et al. (2008b) also reported a near-significant 5-fold increase in the number of apoptotic cells in males exposed to 200 mg/kg/day (p=0.067). Rats exposed to pentaBDEat 200 mg/kg/day for 21or 28 days via gavage showed “minor lesions” in the liver, but no furtherinformation regarding the type or incidence of the lesion was reported (Bruchajzer et al. 2010). In mice, exposure to dietary tetraBDE at 0.45 mg/kg/day for 28 days induced hepatocyte vacuolation, pyknoticnuclei in the hepatocytes, and periportal lymphocytic infiltration; no lesions were observed in controlanimals (Maranghi et al. 2013). In rats exposed to pentaBDE via gavage for 70 days prior to mating until PND 42, significantly increased incidences of hepatocellular hypertrophy were observed in F0 males at ≥5 mg/kg/day and F0 females and F1 offspring at 25 mg/kg/day; no exposure-related effects were observed at 0.5 mg/kg/day (Bondy et al. 2013). Similarly, all F1 rats exposed to pentaBDE at 50 mg/kg/day from GD 6 to PNW 16 via gavage showed hepatocellular hypertrophy characterized by enlarged hepatocytes with an increased amount ofcytoplasm, enlarged nuclei, and pale eosinophilic and granular cytoplasm. Hepatocyte vacuolization wasalso significantly increased in exposed F1 males, but not F1 females (Dunnick et al. 2012). No exposure-related changes were observed in serum chemistry markers (ALT, AST, ALP, LDH,cholesterol, triglycerides, albumin, globulin) in rats exposed to dietary pentaBDE at doses up to 100 mg/kg/day for 13 weeks (WIL Research Laboratories 1984), octaBDE at doses up to 750 mg/kg/dayfor 90 days (IRDC 1977), or a penta-decaBDE mixture (52.1% pentaBDE, 44.2% decaBDE, 0.4%octaBDE) at 20 mg/kg/day for 70 days (Ernest et al. 2012). Similarly, no exposure-related changes were observed in serum ALP, ALT, AST, or gamma-glutamyl transferase (GGT) following exposure to pentaBDE at low doses up to 0.015 mg/kg/day via gavage for 90 days (Daubie et al. 2011). Marginalchanges (<2-fold) were observed in shorter-duration rat gavage studies with higher doses. A statisticallysignificant 51% increase in serum ALT level was observed in male rats exposed to pentaBDE at250 mg/kg/day via gavage for 28 days; no changes were observed in males at 2.5 or 25 mg/kg/day orfemales at any dose (Oberg et al. 2010). No exposure-related changes were observed in serum ALP(Oberg et al. 2010). Similarly, in female rats exposed to pentaBDE at 200 mg/kg/day via gavage for 21 or28 days, serum ALT levels were increased to 190% of control values at 21 or 28 days and serum ASTlevels were increased to 185% of control values at 28 days; no changes were observed at ≤40 mg/kg/day PBDEs 1473. HEALTH EFFECTS(Bruchajzer et al. 2010). Increases in serum cholesterol, total protein, and ALT and decreases in thealbumin/globulin ratio were observed in rats exposed to pentaBDE at 250 mg/kg/day, but the magnitudesof these effects were not reported (Fattore et al. 2001). In another gavage study, rats were exposed to pentaBDE at doses of 0, 0.27, 0.82, 2.47, 7.4, 22.2, 66.7, or 200 mg/kg/day for 28 days (Van der ven et al. 2008b). The study authors reported a dose-related increase in serum ALT in male rats (maximal increase of 148.1%); however, the lowest dose at which the effect was observed was not reported. Instead, resultswere reported in terms of BMD analysis (BMD/BMDLRD10%=61.4/15.5 mg/kg/day). No changes in serum ALT were observed in female rats, and no changes in serum ALP were observed in either sex,although the authors noted procedural problems with their ALP assay (Van der ven et al. 2008b). Thestudy authors reported dose-related increases in serum cholesterol in male and female rats (maximalincreases in males and females were 257 and 144%, respectively). Again, the lowest dose at which the effect was observed was not reported (BMD/BMDLRD10%=15.4/8.4 mg/kg/day in males and 22.4/11.6 mg/kg/day in females) (Van der ven et al. 2008b). In mice, serum ALT levels were significantly elevated by approximately 2-fold in males exposed to tetraBDE at gavage doses of150 mg/kg/day for 12 weeks, compared with control (only dose tested); no other serum biochemistryendpoints were evaluated (Zhang et al. 2015a, 2015b).Significant changes in hepatic oxidative stress parameters were observed in rats following exposure topentaBDE at doses ≥8 mg/kg/day via gavage for 21 or 28 days, including increased hepatic GSH and MDA levels (Bruchajzer et al. 2010). Oxidative stress markers were also significantly altered in the liversof rat offspring following exposure to pentaBDE from GD 6 to PND 21, including significantly increased activities of catalase (CAT) activity at ≥1 mg/kg/day and SOD at 2 mg/kg/day (Blanco et al. 2014). Therewas no change in total levels of thiobarbituric acid reactive substances (TBARS) in offspring (Blanco etal. 2014). In mice, exposure to diBDE at 1.2 mg/kg/day via gavage for 28 days caused significantlydecreased levels of GSH, decreased activities of SOD and glutathione peroxidase (GPx), and increased levels of MDA in the liver (Zhang et al. 2014). Female rats were examined for hepatic porphyria following exposure to pentaBDE or octaBDE at doses of 0, 2, 8, 40, or 200 mg/kg/day via gavage for 21 or 28 days (Bruchajzer 2011; Bruchajzer et al. 2012). PentaBDE caused significant dose-related elevations in total porphyrin levels in the liver followingexposure to ≥8 mg/kg/day for 21 days (~3–8-fold) or 28 days (~3–19-fold), compared to vehicle controls(Bruchajzer 2011), OctaBDE cause significant elevations in liver concentrations of high carboxylatedporphyrins (octa- and hepta-) following exposure to ≥8 mg/kg/day for 21 days (~2–3-fold) or 2 or8 mg/kg/day for 28 days (~3–4-fold) (Bruchajzer et al. 2012). At 28 days, high carboxylated porphyrin PBDEs 1483. HEALTH EFFECTSlevels in the 40 and 200 mg/kg/day groups were not were not significantly elevated compared to the vehicle control. However, the vehicle controls at this duration had an unusually high hepatic porphyrin levels (4-fold increase compared with untreated control). When compared with the untreated controls, porphyrin levels were significantly increased by ~10-, 14-, 4-, and 5-fold in the 2, 8, 40, and200 mg/kg/day groups, respectively. Lower carboxylated porphyrins were <1% of total liver porphyrinsmeasured, and were not further analyzed. These studies also measured ALA-S and ALA-D activity, twoenzymes involved in heme biosynthesis. The results were mixed. PentaBDE exposure caused a significant elevation in ALA-S activity at ≥8 mg/kg/day after 21 days and at ≥2 mg/kg/day after 28 days, while octaBDE exposure caused a significant decrease in ALA-S activity at 200 mg/kg/day after 21 days and at ≥40 mg/kg/day after 28 days. ALA-D activity was also significantly decreased in rats exposed to octaBDE at ≥40 mg/kg/day for 21 or 28 days (ALA-D activity was not assessed in pentaBDE-exposed rats) (Bruchajzer 2011; Bruchajzer et al. 2012). Hepatic porphyria was also observed in rats afterexposure to dietary pentaBDE for 90 days (WIL Research Laboratories 1984). Liver porphyrins weresignificantly elevated by 8.5- and 390-fold in males and females from the 100 mg/kg/day group, respectively, and 3-fold in females from the 20 mg/kg/day group; no significant changes were observed in the 2 mg/kg/day group (WIL Research Laboratories 1984). Elevated liver weights have been reported following intermediate-duration exposure to penta-, octa-, tetra-, or diBDE in several animal studies. Significantly increased liver weights were reported in rats andmice exposed to penta-, di-, or tetraBDE via gavage at ≥1.2, ≥50, and 150 mg/kg/day, respectively, for15–84 days (Becker et al. 2014; Bruchajzer et al. 2010; Fattore et al. 2001; Mercado-Feliciano and Bigsby2008a; Oberg et al. 2010; Stoker et al. 2004, 2005; Van der Ven et al. 2008b; Zhang et al. 2014, 2015a, 2015b). No exposure-related changes in liver weight were observed in mice exposed to tetraBDE at1 mg/kg/day via gavage for 6 weeks (only dose tested) (McIntyre et al. 2015). In dietary studies, significantly increased liver weights were reported in rats exposed to penta- or octaBDE at ≥5 mg/kg/day for 28–90 days (IRDC 1976, 1977; WIL Research Laboratories 1984) or a penta-decaBDE mixture(52.1% pentaBDE, 44.2% decaBDE, 0.4% octaBDE) at 20 mg/kg/day for 70 days (Ernest et al. 2012). In mink exposed to dietary pentaBDE at 0, 0.08, 0.63, or 0.78 mg/kg/day for 90 days, increased absoluteliver weight was observed at 0.78 mg/kg/day and increased relative liver weight was observed at≥0.08 mg/kg/day; however, these findings are confounded by significant body weight loss at≥0.63 mg/kg/day (Martin et al. 2007). No exposure-related changes in liver weight were observed inmice exposed to dietary tetraBDE at 0.45 mg/kg/day for 28 days (Maranghi et al. 2013). PBDEs 1493. HEALTH EFFECTSIn studies that evaluated hepatic enzyme induction (e.g., elevated EROD, MROD, PROD, and UDPGTactivity), significantly increased enzyme activities were observed in rats, mice, and mink at doses of ≥2.5, ≥50, and ≥0.08 mg/kg/day, respectively, and were always observed at doses at or below the dose causingelevated liver weights in the same study (Bruchajzer et al. 2010; Ernest et al. 2012; Fattore et al. 2001;Martin et al. 2007; Mercado-Feliciano and Bigsby 2008a; Oberg et al. 2010; Stoker et al. 2004). Additionally, microsomal enzyme activity was induced in rats exposed by gavage to doses as low as 0.6 mg/kg/day of octaBDE and 0.4 mg/kg/day of pentaBDE for 90 days as indicated by increases inO-ethyl O-p-nitrophenyl phenylphosphonothioate (EPN) detoxification, p-nitroanisole demethylation, and cytochrome c reductase and cytochrome P-450 levels (Carlson 1980a). Some of these changes werepersistent, lasting for 30–60 days after cessation of treatment.Maternal and pup liver weights were significantly elevated in rats exposed to pentaBDE from GD 1 orGD 6 to PND 21 via gavage or dosed cookies at 30 and ≥3 mg/kg/day, respectively (Bowers et al. 2015;Ellis-Hutchings et al. 2006; Zhou et al. 2002). In mice, elevated maternal liver weights were observed indams exposed to 452 mg/kg/day of the pentaBDE congener BDE 99, but not the pentaBDE commercialmixture Bromkal 70-5DE, from GD 6 to PND 17 via gavage (Skarman et al. 2005). Elevated liverweights were not observed in mouse pups (Skarman et al. 2005). In a one-generation study in rats (premating day 70 to PND 42), F0 males and F1 offspring showed significantly increased liver weights on PND 43 following exposure to 25 and ≥5 mg/kg/day via gavage (Bondy et al. 2011, 2013). In other one-generation studies, no changes were observed in maternal or pup liver weight in rats exposed topentaBDE-dose vanilla wafers at doses up to 11.4 mg/kg/day from pre-mating day 28 to PND 21 (Poon etal. 2011) or adult F1 liver weight in rats exposed to pentaBDE at 50 mg/kg/day from GD 6 to PNW 16 via gavage (Dunnick et al. 2012). In mink, exposure to dietary pentaBDE from 4 weeks premating untilPNW 6 or 33 led to elevated liver weights in sows at 0.31 mg/kg/day (highest dose tested) and F1 offspring at 0.06 mg/kg/day (Zhang et al. 2009). In the studies that evaluated hepatic enzyme levels (e.g., EROD, PROD, UDPGT), significantly increased enzyme activities were always observed at doses at or below the dose causing elevated liver weights in the same study, with significant changes observedin F0 and F1 rats at ≥0.3 mg/kg/day, F0 and F1 mice at 450 mg/kg/day, and F0 and F1 mink at ≥0.06 and 0.01 mg/kg/day, respectively (Bowers et al. 2015; Skarman et al. 2005; Zhang et al. 2009; Zhou et al. 2002). Additionally, significant induction of hepatic enzymes was observed in male rat pups followingexposure to pentaBDE at ≥1.7 mg/kg/day from GD 6 to PND 21 via gavage (Szabo et al. 2009). Szabo et al. (2009) also evaluated genomic changes in F1 rats following exposure to pentaBDE at 0, 1.7, 10.2, or 30.6 mg/kg/day from GD 6 to PND 4 or 21 via gavage. Significant dose-related increases were PBDEs 1503. HEALTH EFFECTSobserved in hepatic mRNA expression levels of phase I enzymes (Cyp1a1, Cyp2b1, Cyp2b2, Cyp3a1), phase II enzymes (Ugt1a6, Ugt1a7, Ugt2b, Sult1b1), phase III influx transporters (Oatp1a4), phase IIefflux transporters (Mdr1, Mrp2, Mrp3), the serum binding protein transthyretin (Ttr), and deiodinase 1 (d1) during exposure (PND 4 and 21); the majority of which no longer differed from control at PND 60. Genomic changes were also evaluated in F1 rats following exposure to pentaBDE at 0, 1, or 2 mg/kg/dayvia gavage (Blanco et al. 2014). Various cytochrome isoforms were evaluated, but mRNA levels were only significantly elevated for CYP2B1 in pup liver tissue from the 2 mg/kg/day group. Additionally, significant changes were observed in the thyroid hormone receptor, including decreased TRα1 mRNA in both dose groups, decreased TRβ1 mRNA at 2 mg/kg/day, and decreased TRα1 protein levels at2 mg/kg/day. Cylcin D1 protein expression was significantly decreased in both groups and thephosphorylation levels of Akt and GSK3β were significantly decreased at 2 mg/kg/day.To determine if PBDE exposure alters vitamin A homeostasis in a manner similar to the related PBBs(ATSDR 2004), hepatic vitamin A levels were measured in rats following exposure to pentaBDE at 0, 2.5, 25, or 250 mg/kg/day via gavage for 28 days (Oberg et al. 2010). Hepatic liver vitamin A content was significantly decreased in females at 25 mg/kg/day and males and females at 250 mg/kg/day by up to 36 and 47%, respectively (Oberg et al. 2010). Hepatic vitamin A levels were also significantly decreasedby 13% in dams and 50% in F1 pups on PND 18 in rats exposed to pentaBDE at 0 or 18 mg/kg/day fromGD 6 to PND 18 via gavage (Ellis-Hutchings et al. 2006). Additional pup sacrifices on PNDs 12 and 31 showed significant 59 and 25% decreases in hepatic vitamin A level, respectively (Ellis-Hutchings et al. 2006). DecaBDE: Moderate hepatocellular hypertrophy was observed in all mice exposed to dietary decaBDE at9,400 mg/kg/day for 28 days, but in none of the controls (Sakamoto et al. 2013). Similarly, slightcentrilobular hypertrophy was “occasionally” observed in rats exposed to decaBDE at doses ranging from 1.87 to 60 mg/kg/day for 28 days via gavage, which was most obvious in “some” of the 60 mg/kg/daymales (incidence data not reported) (Van der ven et al. 2008a). In an older study using an impuredecaBDE compound (77% purity), centrilobular cytoplasmic enlargement and vacuolation were observed in male rats exposed to dietary doses of 800 mg/kg/day for 30 days (incidences not reported); no changeswere observed at 8 or 80 mg/kg/day (Norris et al. 1973, 1975a). However, no exposure-related changes in liver histology were observed in rats and mice exposed to dietary decaBDE at estimated doses as highas 2,000–8,000 and 2,375–9,500 mg/kg/day for 13 weeks, respectively (NTP 1986), or in rats exposed to gavage doses up to 1,000 mg/kg/day for 21 or 28 days (Bruchajzer et al. 2010). PBDEs 1513. HEALTH EFFECTSPregnant dams and developing rats and mice appear to be more susceptible to liver damage followingexposure to decaBDE than adult animals. Hepatocyte degeneration and eosinophil changes wereobserved in the livers of rat dams exposed to decaBDE at 300 mg/kg/day for 21 days prior to mating, through mating, gestation, and lactation (PND 21) (Liu et al. 2012). Compared with control, the exposedgroup had significantly increased “histological scores;” however, the method of histological scoring andincidences of lesions were not reported (Liu et al. 2012). Following exposure to dietary decaBDE fromGD 10 to PND 21, male and female rat pups showed significantly increased incidence of follicular cellhypertrophy in the liver at ≥2 and 146 mg/kg/day, respectively (Fujimoto et al. 2011). These findingswere transient, as they were no longer evident in male and female offspring evaluated at PNW 11(Fujimoto et al. 2011). Male offspring from mouse dams exposed to decaBDE from GD 0 to 17 via gavage showed dose-related histopathological changes in the liver on PND 71 at doses ≥10 mg/kg/day(lowest dose tested) (Tseng et al. 2008). Histological changes observed in exposed groups included acutecell swelling of hepatocytes associated with pressure occlusion of hepatic sinusoids (Tseng et al. 2008). Young male rats exposed to decaBDE at doses of 0, 100, 300, or 600 mg/kg/day from PND 10 to 42 viagavage showed fatty degeneration at ≥300 mg/kg/day (incidence data were not reported) (Lee et al. 2010).Wang et al. (2010) found no exposure-related changes in serum ALT, AST, or ALP in male rats followinggavage exposure to decaBDE at 100 mg/kg/day for 90 days, but serum total cholesterol, high densitylipid-cholesterol, and total bile acid levels were significantly increased by 23, 26, and 98%, respectively. No biologically relevant, exposure-related changes were observed in either serum ALT or AST, or totalcholesterol or triglyceride levels, in rats exposed to decaBDE at doses up to 1,000 mg/kg/day for 21– 56 days (Bruchazjer et al. 2010; Van der ven et al. 2008a; Zhang et al. 2013a).No exposure-related changes were observed in hepatic markers of oxidative stress (GSH, MDA) infemale rats exposed to decaBDE up to doses of 1,000 mg/kg/day for 21 or 28 days (Bruchazjer et al. 2010).Significantly elevated liver weights were observed in male rats after exposure to decaBDE at≥1 mg/kg/day for 8 weeks via gavage (Zhang et al. 2013a) and male mice after exposure to dietarydecaBDE at 9,400 mg/kg/day for 28 days (Sakamoto et al. 2013). However, no exposure-related changes in liver weight were observed in rats exposed to decaBDE up to doses of 1,000 mg/kg/day via gavage for21–90 days (Bruchazjer et al. 2010; Van der ven et al. 2008a; Wang et al. 2010, 2011b). In an older studyusing an impure decaBDE compound (77% purity), increased liver weights were observed at dietarydoses ≥80 mg/kg/day in male rats exposed for 30 days (Norris et al. 1973, 1975a). Following exposure to PBDEs 1523. HEALTH EFFECTSdietary decaBDE from GD 10 to PND 21, male and female rat pups showed significantly increased liverweights at ≥2 and 146 mg/kg/day, respectively (Fujimoto et al. 2011). These findings were transient, as they were no longer evident in male and female offspring evaluated at PNW 11 (Fujimoto et al. 2011).No liver weight changes were observed in dams or PND 71 male offspring exposed to decaBDE fromGD 0 to 17 via gavage at doses up to 1,500 mg/kg/day (Tseng et al. 2008). Young male rats exposed to decaBDE at doses of 0, 100, 300, or 600 mg/kg/day from PND 10 to 42 via gavage showed elevated liverweights at ≥300 mg/kg/day (Lee et al. 2010).Hepatic enzyme induction (e.g., EROD, PRO activity, CYP protein levels) was significantly elevated in adose-related manner in adult male and female rats exposed to decaBDE at 1.87–60 mg/kg/day for 28 days via gavage (Van der ven et al. 2008a) and young male rats exposed to decaBDE at 100–600 mg/kg/day from PND 10 to 42 via gavage (Lee et al. 2010). EROD activity was also significantly elevated inPND 71 male mice exposed to 1,500 mg/kg/day during gestation (Tseng et al. 2008). However, no exposure-related changes were observed in hepatic enzyme induction in female rats exposed to decaBDEup to doses of 1,000 mg/kg/day for 21 or 28 days (Bruchazjer et al. 2010). Genomic analyses of livertissue reported elevated mRNA levels of CYP proteins in male mice exposed to dietary decaBDE at9,400 mg/kg/day for 28 days (Sakamoto et al. 2013) and a dose-dependent elevation of Cyp2b2 mRNA inmale rats exposed to decaBDE at doses of 1.87–60 mg/kg/day for 28 days via gavage (Van der ven et al. 2008a).To determine if PBDE exposure alters vitamin A homeostasis in a manner similar to the related PBBs(ATSDR 2004), hepatic vitamin A levels were measured in rats following exposure to decaBDE at 0, 1.87, 3.75, 7.5, 15, 30, or 60 mg/kg/day for 28 days via gavage (Van der ven et al. 2008a). In females, but not males, hepatic vitamin A levels were increased in a dose-dependent manner by up to 14.6%, compared with controls; however, the lowest dose at which the effect was observed was not reported.The study authors conclude that the relevance of the effect is uncertain due to high variation in the data, as evidenced by a high BMD/BMDL ratio (BMD/BMDLRD10% =13.8/1.2 mg/kg/day).Chronic-Duration Animal StudiesLower-brominated PBDEs: No chronic-duration studies analyzing hepatic effects were located for lowerbrominated PBDEs. PBDEs 1533. HEALTH EFFECTSDecaBDE: In chronic studies, exposure to 94–97% decaBDE for 103 weeks caused liver lesions that included neoplastic nodules in rats at ≥1,120 mg/kg/day, thrombosis and degeneration in rats at2,240 mg/kg/day, and centrilobular hypertrophy and granulomas in mice at ≥3,200 mg/kg/day (NTP1986). The thrombosis in the rats was characterized by a near total occlusion of a major hepatic blood vessel by a dense fibrin coagulum. A NOAEL was not identified in the rats or mice. The only otherchronic study of decaBDE found that exposure to 1 mg/kg/day of a 77% pure mixture for 2 years causedno liver effects in rats; higher doses were not tested, precluding identification of a LOAEL (Kociba et al.1975; Norris et al. 1975a).Summary. No studies are available on hepatic effects of PBDEs in humans. Based on the evidence inanimals, including enzyme induction, liver enlargement, histological lesions, hepatic porphyria, and markers of hepatic oxidative stress, PBDEs are potentially hepatotoxic in humans, especially lowerbrominated PBDEs.Renal Effects. Only one study was located that provides information on renal effects in humans following exposure to PBDEs. A pilot study measured serum creatinine and urinary β2-microglobulin, asindices of renal function, in 40 Chinese residents exposed to PBDEs in an electronic waste dismantlingarea (Xu et al. 2015a). Mean ΣPBDEs in serum from exposed subjects was almost twice that in a controlgroup, but the difference was not statistically significant. The results showed that neither serumcreatinine nor β2-microglobulin were correlated with PBDEs. However, congener analyses showed thatBDE 28, BDE 47, BDE 85, and BDE 153 were positively correlated with urinary levels ofβ2-microglobulin, and that BDE 28 and BDE 85 were positively correlated with serum creatinine levels. PCBs also played a role in the findings reported.Acute-Duration Animal StudiesLower-brominated PBDEs: Two studies examined renal effects in male rats 45 days after a single gavage administration of pentaBDE at doses of 0, 0.06, or 1.2 mg/kg (Albina et al. 2010; Alonso et al. 2010). In urine, a 4-fold increase in total protein levels was observed at 1.2 mg/kg (Alonso et al. 2010). Although this might suggest possible tubular damage, histopathological examination of the kidneys found no abnormalities other than a dose-related increase in phagolysosomes (incidence data not reported) (Albinaet al. 2010). No other changes were observed in urinalysis or serum chemistry parameters (urea, creatinine, uric acid) (Alonso et al. 2010). Altered oxidative stress markers were found in the kidney after PBDEs 1543. HEALTH EFFECTSexposure to 1.2 mg/kg/day, including significantly decreased CAT activity and increased GSSG and GSSG/GSH ratio (Albina et al. 2010). No exposure-related changes in kidney weight were observed in male rats exposed to pentaBDE at dosesup to 240 mg/kg/day for 9 days via gavage (Stoker et al. 2005) or female mice exposed to tetraBDE atdoses up to 100 mg/kg/day for 4 days via gavage (Richardson et al. 2008). DecaBDE: No acute-duration studies analyzing renal effects were located for decaBDE.Intermediate-Duration Animal StudiesLower-brominated PBDEs: Renal effects induced by dietary octaBDE included non-inflammatorykidney changes in male rats exposed to 600 mg/kg/day for 13 weeks, but not females exposed to doses upto 750 mg/kg/day (IRDC 1977). The incidence and severity of the kidney lesions (tubule regeneration, intratubular casts, and cellular debris occurred in most 600 mg/kg/day males) suggested a compound-related effect (IRDC 1977). However, no histopathological lesions of the kidney were observed in maleor female rats exposed to pentaBDE at doses up to 250 mg/kg/day for 28 days via gavage (Oberg et al. 2010; Van der ven et al. 2008b), dietary penta- or octaBDE at doses up to 90 mg/kg/day for 28 days(IRDC 1976), or dietary pentaBDE doses up to 100 mg/kg/day for 90 days (WIL Research Laboratories 1984). Additionally, no histopathological kidney lesions were observed in F0 or F1 rats exposed to pentaBDE at doses up to 25 mg/kg/day from pre-mating through PND 42 (Bondy et al. 2013).Statistically significant changes in blood urea and urea nitrogen levels were reported in some intermediate-duration studies; however, none of the changes were considered biologically relevant due tothe small magnitude of change (<2-fold) compared with control. Following exposure to pentaBDE at 0, 2.5, 25, or 250 mg/kg/day via gavage for 28 days, blood urea levels in male rats from the 25 or250 mg/kg/day groups were significantly increased by 1.5- and 1.2-fold, respectively; no changes were observed in blood urea levels in females or blood creatinine in either sex (Oberg et al. 2010). In another28-day study, rats were exposed to pentaBDE at 0, 0.27, 0.82, 2.47, 7.4, 22.2, 66.7, or 200 mg/kg/day viagavage (Van der ven et al. 2008b). Study authors reported dose-related increases in blood urea levels inmale and female rats (maximal increases in males and females were 61.2 and 33.1%, respectively);however, the lowest dose at which the effect was observed was not reported. Instead, results werereported in terms of BMD analysis (BMD/BMDLRD10%=64.2/30.2 mg/kg/day in males and 65.1/22.2 mg/kg/day in females) (Van der ven et al. 2008b). Again, no changes in serum creatinine were PBDEs 1553. HEALTH EFFECTSreported in either sex at doses up to 200 mg/kg/day (Van der ven et al. 2008b). Blood urea nitrogen (BUN) levels were elevated by 35% in male rats and 59% in female rats exposed to dietary octaBDE for13 weeks at 600 and 750 mg/kg/day, respectively (IRDC 1977). No changes in BUN levels wereobserved in male or female rats exposed to dietary pentaBDE doses up to 100 mg/kg/day for 90 days(WIL Research Laboratories 1984). Kidney weight changes following intermediate-duration exposure are inconsistent. A significant 10%decrease in relative kidney weight was observed in mice exposed to diBDE at 1.2 mg/kg/day for 28 daysvia gavage (Zhang et al. 2014). Absolute organ weights were not reported; however, no body weighteffects were observed. In contrast, no change in kidney weight was reported in rats following exposure to pentaBDE at gavage doses up to 120 mg/kg/day for 20–90 days (Daubie et al. 2011; Stoker et al. 2004, 2005) or dietary penta- or octaBDE at doses up to 750 mg/kg/day for 28–90 days (IRDC 1976, 1977;WIL Research Laboratories 1984). Similarly, no change in F0 or F1 kidney weight was observed in ratsexposed to pentaBDE at doses up to 25 mg/kg/day via gavage from 70 days prior to mating through PND 42 (Bondy et al. 2013) or mink exposed to dietary pentaBDE at doses up to 0.31 mg/kg/day from4 weeks prior to mating to PNW 6 or 33 (Zhang et al. 2009). In other studies, significantly increased kidney weights have been reported, including a dose-related increase in absolute liver weight in rats exposed to pentaBDE doses up to 200 mg/kg/day for 28 days via gavage (dose at which effect was firstobserved was not reported; maximum increase of 11.5% in males and 15.6% in females) (Van der ven etal. 2008b), a significant 15% increase in relative, but not absolute, kidney weight following exposure to pentaBDE at 250 mg/kg/day for 28 days via gavage (Oberg et al. 2010), and a significant 18% increase inrelative kidney weight following exposure to a dietary penta-decaBDE mixture (52.1% pentaBDE, 44.2%decaBDE, 0.4% octaBDE) at 20 mg/kg/day for 70 days (absolute kidney weight not reported) (Ernest etal. 2012).To determine if PBDE exposure alters vitamin A homeostasis in rats in a manner similar to the relatedPBBs (ATSDR 2004), vitamin A levels were measured in kidney tissue following exposure to pentaBDEat 0, 2.5, 25, or 250 mg/kg/day via gavage (Oberg et al. 2010). No significant changes in kidney vitamin A levels were observed at any dose.DecaBDE: No renal histopathological changes were observed in rats or mice exposed to dietarydecaBDE at doses up to 8,000 or 9,500 mg/kg/day, respectively, for 13 weeks (NTP 1986). A 28-day study that evaluated histology of “major organs” did not report any exposure-related kidney effects in rats exposed to decaBDE at doses up to 60 mg/kg/day via gavage (Van der ven et al. 2008a). Studies of low PBDEs 1563. HEALTH EFFECTSpurity (≈77%) commercial decaBDE mixtures found kidney pathology (hyaline degenerative cytoplasmicchanges) in male rats exposed to dietary levels of 800 mg/kg/day for 30 days (Norris et al. 1973,1975a), but not in rats exposed to ≤90 mg/kg/day for 28–30 days (IRDC 1976; Norris et al. 1973, 1975a). Interpretation of this finding is complicated by the fact that hyaline degenerative cytoplasmic changes are not uncommon in adult male rats and might be induced by a mechanism specific to certain aged male rats.No exposure-related changes were observed in serum urea or creatinine levels in rats exposed todecaBDE at doses up to 100 mg/kg/day for 28–90 days via gavage (Van der ven et al. 2008a; Wang et al.2010). Additionally, no exposure-related changes were observed in kidney weights in rats exposed todecaBDE at doses up to 600 mg/kg/day for 28–90 days via gavage (Lee et al. 2010; Van der ven et al. 2008a; Wang et al. 2010, 2011b; Zhang et al. 2013b), F0 or F1 mice exposed to decaBDE at doses up to 1,500 mg/kg/day from GD 0 to 17 via gavage (Tseng et al. 2008), or dietary decaBDE at doses up to 800 mg/kg/day for 28–30 days (IRDC 1976; Norris et al. 1973, 1975a).Chronic-Duration Animal StudiesLower-brominated PBDEs: No chronic-duration studies analyzing renal effects were located for lowerbrominated PBDEs.DecaBDE: No renal histopathological changes were observed in rats or mice exposed to dietarydecaBDE at doses up to 2,550 or 7,780 mg/kg/day, respectively, for 2 years (NTP 1986). The only otherchronic study of decaBDE found that exposure to dietary doses up to 1 mg/kg/day of the 77% puremixture for 2 years caused no exposure-related changes in kidney histology or weight in rats (Kociba etal. 1975; Norris et al. 1975a). Summary. No studies are available on hepatic effects of PBDEs in humans. While there is limitedevidence from animal studies that lower-brominated PBDEs can cause kidney damage at high exposure levels, data are inconsistent and there is no evidence of impaired renal function. Animal studies do notindicate that decaBDE causes renal toxicity. Taken together, animals studies indicate that renal effectsare not likely to occur in humans at environmentally-relevant exposure concentrations. PBDEs 1573. HEALTH EFFECTSEndocrine Effects.Human Studies. Numerous studies have been performed to evaluate the relationship between concentrations of PBDE in body tissues and circulating hormone levels in human populations. Whilethese studies have demonstrated that PBDE can perturb the human endocrine system and affect hormonelevels, the specific findings are not consistent across studies. For example, even limiting the discussion tostudies that evaluated both PBDE concentrations and thyroid hormone levels in serum samples collected only from adult men, studies have reported positive associations with T4 (Turyk et al. 2008), negativeassociations with T4 (Abdelouahab et al. 2011), and no association with T4 (Hagmar et al. 2001). Thesestudies also reported either negative association with T3 (Turyk et al. 2008) or no association with T3 (Abdelouahab et al. 2011; Hagmar et al. 2001), and either negative association with TSH (Hagmar et al.2001; Turyk et al. 2008) or no association with TSH (Abdelouahab et al. 2011). Populations in thesestudies were 308 adult male consumers of sport fish from U.S. Great Lakes with serum ∑PBDE rangingfrom 15.8 to 1,360 ng/g lipid (0.13–10.15 ng/g) with a median of 38.4 ng/g lipid (0.26 ng/g) (Turyk et al. 2008), 48 adult men recruited through an infertility clinic in Quebec with median serum ∑PBDE of0.302 ng/mL and maximum of 2.250 ng/mL (Abdelouahab et al. 2011), and 110 adult men from Latviaand Sweden having varying consumption of Baltic Sea fish with serum BDE 47 ranging from 0.10 ng/g lipid (10th percentile) to 5.16 ng/g lipid (90th percentile) and a median of 1.04 ng/g lipid (Hagmar et al. 2001).Similarly, studies of serum PBDE and serum thyroid hormones in populations including adults of both sexes found: (1) a negative association with free T4 and no association with T4, T3, or TSH in 114 elderlyresidents of the upper Hudson River area of New York State (∑PBDE range of 0.04–9.80 µg/L and median of 0.19 µg/L in 48 women; range of 0.04–4.74 µg/L and median of 0.16 µg/L in 66 men) (Bloomet al. 2014); (2) a negative association with total T4 (BDE 47, BDE 99, and BDE 100) and a positiveassociation with TSH (BDE 153 only), but no association with free T4 or total T3, in a longitudinal studyof 51 healthy adult office workers from Boston with serum ∑PBDE geometric means ranging from 19 to 23 ng/g lipid over 18 months (Makey et al. 2016); (3) no association with free or total T4, T3, or TSH in 36 New York anglers living in counties adjacent to Lakes Erie or Ontario with median serum ∑PBDE of15 ng/g lipid and maximum of 2,303 ng/g lipid (Bloom et al. 2008); (4) a positive association with T3 (for BDE 47 only) and no association with free T4 or TSH in 623 Nunavik Inuits with geometric mean BDE47=2.16 and BDE 153=2.05 µg/kg lipid (Dallaire et al. 2009); (5) no association with T4, T3, or TSH in 11 Swedish electronics recycling workers sampled repeatedly over 1.5 years (∑PBDE median=7.2 pmol/glipid weight at start of employment and 9.7 pmol/g lipid weight at the conclusion of the study) (Julander PBDEs 1583. HEALTH EFFECTSet al. 2005); (6) a positive association with TSH in 23 Chinese e-waste workers (∑PBDEmedian=382 ng/g lipid with range of 77–8,452 ng/g lipid) versus 26 controls (∑PBDE median=158 ng/glipid with range of 18–436 ng/g lipid) (Yuan et al. 2008); (7) a positive association with T4 (for BDE 126 and BDE 205 only) and no association with free T4, T3, free T3, or TSH in another group of 239 Chinesee-waste workers (∑PBDE median=189.79 ng/g lipid with range of 0–6,016 ng/g lipid) and 93 farmersfrom the e-waste area (∑PBDE median=164.64 ng/g lipid with range of 0–8,600 ng/g lipid) versus116 controls (∑PBDE median=122.37 ng/g lipid with range of 0–1,398 ng/g lipid) (Wang et al. 2010); (8) a negative association with T3 (for BDE 17 and BDE 153 only) and no association with free T4 orTSH in 124 residents of northern China with serum ∑PBDE median=7.16 ng/g lipid with range of 2.09– 160.3 ng/g lipid (Huang et al. 2014); (9) no associations between serum PBDEs or their methyoxylated orhydroxylated metabolites and thyroid hormones in 77 residents living near an e-waste recycling site inVui Dau, Vietnam or 34 residents living in rural Doung Quang, Vietnam (median ∑PBDE serumconcentrations of 290 and 230 pg/g wet weight, respectively) (Eguchi et al. 2015); and (10) no associations between serum PBDEs and thyroid hormones in 40 residents living near an e-waste recyclingsite in Luqiao, China or 15 residents living in rural Yunhe, China (median ∑PBDE serum concentrationsof 51.61 and 66.45 ng/g, respectively) (Xu et al. 2015a). Similar studies performed in children found:(1) a positive association with serum PBDE for serum TSH in 195 6–8-year-old children from an e-waste recycling area in China (∑PBDE mean=664.28 ng/g lipid) or 174 children from a control area (∑PBDEmean=375.81 ng/g lipid) (Han et al. 2011); (2) a negative association with serum PBDE for free T3 and apositive association with serum PBDE for TSH in 162 4–6-year-old children living in an e-waste area inChina (serum ∑PBDE median=189.99 ng/g lipid) (Xu et al. 2014b); (3) no association between free and total T3 and T4, TSH, and serum PBDE in 21 8-year-old children from an e-waste recycling area in China (∑PBDE median=31.86 ng/g lipid) or 24 children from a control area (∑PBDE median=6.97 ng/g lipid)(Xu et al.2014a); (4) positive associations with T3 and free T4 (for BDE 99 only) and no association with T4 or TSH in 17 Dutch teenagers with serum ∑PBDE ranging from 4.9–22.1 ng/g lipid and a mean of10.5 ng/g lipid (Leijs et al. 2012); and (5) and negative or no association with free T3 (depending on typeof analysis), positive or no association with TSH, and no association with free T4 in 515 Flemish teenagers with a median serum ∑PBDE of 7 ng/L and maximum of 125 ng/L (Kicinski et al. 2012). Oulhote et al. (2016) examined the potential association between PBDE exposure and hypothyroidism in Canadian women aged 30–79 years. PBDE levels were not significantly different in women diagnosedwith hypothyroidism (n=90; geometric mean serum ∑PBDE=15.4 ng/g lipid) compared with women without hypothyroidism (n=655; geometric mean serum ∑PBDE=20.5 ng/g lipid). In a model adjustedfor age, income, education, alcohol consumption, race/ethnicity, and history of breast-feeding, the PBDEs 1593. HEALTH EFFECTSprevalence ratio of hypothyroidism was marginally increased by 1.7 per 10-fold increase in ∑PBDEs (95% CI 1.0, 3.0) or BDE 47 (95% CI 1.0, 2.9) and marginally increased by 1.8 in individuals with detectable BDE 100 compared with those without detectable BDE 100 (95% CI 1.0, 3.3). When modelswere adjusted for age income, education, alcohol consumption, and lipid-standardized PCB-153, thesemarginal associations were no longer observed.A couple of studies were located in which exposure was estimated from concentrations of PBDE in housedust, rather than serum samples. In the earlier study, Meeker et al. (2009) found a positive association between PBDE in house dust and serum levels of free T4, but no association with T3 or TSH in 24 men recruited through an infertility clinic. Median and maximum concentrations of PBDE in dust collected from houses of study subjects were 500 and 7,620 ng/g dust for BDE 47, 838,and 9,220 ng/g dust for BDE 99, and 180 and 2,830 ng/g dust for BDE 100. A later study by these same researchers found positive associations between dust concentrations of pentaBDE (sum of BDE 47, BDE 99, and BDE 100;median and maximum concentrations of 1,049 and 22,300 ng/g in dust) and octaBDE (sum of BDE 183 and BDE 201; median and maximum concentrations of 30.5 and 1,181 ng/g in dust) and serum free T4, and also between pentaBDE and T3 and octaBDE and TSH in 38 additional men recruited through theinfertility clinic (Johnson et al. 2013). There were no significant associations between thyroid hormonesand exposure to decaBDE (sum of BDE 206, BDE 207, BDE 208, and BDE 209; median and maximum concentrations of 1,800 and 38,483 ng/g in dust) in this study.As in other populations, findings in pregnant women differed across studies. Both free and total T4 were significantly and positively associated with PBDE (several individual congeners and their sum) in serumcollected during the 3rd trimester (>34 weeks) of pregnancy in a North Carolina cohort of 137 primarilyAfrican-American expectant mothers with serum ∑PBDE ranging from 3.59 to 693.95 ng/g lipid and amedian of 36.56 ng/g lipid (Stapleton et al. 2011). Associations for free and total T3 and TSH were notsignificant in this cohort. In contrast, a study of a California cohort of 270 mostly Latina women tested atthe 27th week of pregnancy (serum ∑PBDE range of 3.6–1338.6 ng/g lipid, median of 25.2 ng/g lipid)found no association between PBDE and free or total T4 (Chevrier et al. 2010). However, all PBDEcongeners identified and their sum were significantly negatively associated with TSH in this study.Further analysis showed that women in the highest quartile of PBDE exposure had significantly increased odds of subclinical hyperthyroidism (defined as low TSH and normal free T4) relative to women in thefirst quartile. A study of 105 pregnant women in South Korean that looked at blood samples collected the day before delivery (serum ∑PBDE median=2.13 ng/g lipid with 25th–75th percentile range of 1.35– 4.34 ng/g lipid) found significant negative associations for PBDE with free and total T3, a significant PBDEs 1603. HEALTH EFFECTSpositive association with free T4, and no association with T4 or TSH (Kim et al. 2013a). There were no correlations between PBDE and free and total T4 and free and total T3 in maternal serum samples (∑PBDE range from 15 to 580 ng/g lipid, median=37 ng/g lipid) collected just prior to delivery in anotherstudy of 12 patients from Indiana (Mazdai et al. 2003).Abdelouahab et al. (2013) compared maternal serum concentrations of PBDE collected at <20 weeks ofpregnancy (for ∑PBDE, median=30.92 ng/g lipid, maximum=726.09 ng/g lipid, n=380) with thyroid hormone levels in maternal blood at <20 weeks of pregnancy and in maternal blood collected at deliveryfor 260 pregnancies in Quebec. At <20 weeks of pregnancy, they found significant negative associationsbetween PBDE and T3 and T4, but significant positive associations between PBDE and free T3 and freeT4. However, using the thyroid hormone levels from the maternal samples at delivery, they found negative associations between maternal serum PBDE at <20 weeks and maternal free and total T3 and T4 at delivery. There was no significant association with TSH for either of the samples. In a Swedish cohort(n=166), a significant inverse relationship was observed between first trimester total T3 levels andmaternal body burden of BDE 153 (median breast milk concentration, 0.48 ng/g lipid); this relationship was not significant for third trimester total T3 levels (Lignell et al. 2016). No significant associationswere observed between breast milk BDE 153 levels and free T4 or TSH in first or third trimester serumsamples. No associations were observed between serum thyroid hormone levels and breast milkconcentrations of tetra-pentaBDE (BDE 47, BDE 99, BDE 100; median concentration, 2.3 ng/g lipid).A few studies included analysis of maternal serum samples collected after delivery. Kim et al. (2011d,2012b) found significant positive relationships between serum PBDE (BDE 49 only) and free T4, and between PBDE (BDE 154 and BDE 153 only) and TSH, and a significant negative relationship between PBDE (BDE 153 only) and T3 in maternal blood samples collected from 12 South Korean mothers afterdelivery (for ∑PBDE, range=1.88–53.54 ng/g lipid, mean=18.79 ng/g lipid). There were no such correlations between serum PBDE and thyroid hormone levels in post-delivery blood samples collectedfrom 26 mothers of infants born with congenital hypothyroidism (for ∑PBDE, range=3.81–1563 ng/glipid, mean=65.16 ng/g lipid) (Kim et al. 2011d, 2012b). In a second study in which maternal blood samples were collected after delivery (for 21 South Korean mothers undergoing Cesarean section), there was no correlation between PBDE concentrations and thyroid hormone (free T4, T3, and TSH) levels in the maternal serum (Kim et al. 2012a). In this study, ∑PBDE in maternal blood ranged from 1.8 to 17.66 ng/g lipid, with a median of 7.81 ng/g lipid. Kim et al. (2011a) reported a significant positive correlation between concentrations of BDE 153 in breast milk (mean ≈0.25 ng/g lipid) and serum TSHcollected post-delivery in another group of South Korean mothers. PBDEs 1613. HEALTH EFFECTSIn addition to the maternal effects, several of these studies and others reported on thyroid hormonechanges in developing offspring, as identified by umbilical cord or neonatal blood samples in relation to PBDE concentrations in maternal serum, cord serum, or breast milk. The data on thyroid hormone effectsin developing offspring are presented in Section 3.2.2.6 on Developmental Effects.Few human data were located on endocrine end points other than thyroid (discussed above) or sexhormones (discussed in Section 3.2.2.5, Reproductive Effects). Lim et al. (2008) performed an analysisof cross-sectional U.S. NHANES 2003–2004 data that showed significant increases in risk of diabetesand metabolic syndrome associated with serum concentrations of BDE 153 (but not BDE 28, BDE 47, BDE 99, or BDE 100) based on 156 and 237 cases, respectively, in a population of 1,367 adults examined for diabetes status and, for metabolic syndrome, a subset of 637 participants with a morning fasting blood sample. In both analyses, the risk of disease was higher with exposure to 25th–50th percentile BDE 153 concentrations (median=3.6 ng/g lipid) and 50–75th percentile BDE 153 concentrations (median=6.6 ng/glipid) than >75th percentile BDE 153 concentrations (median=24.6 ng/g lipid). Adjusted ORs were 2.6, 2.7, and 1.8 for diabetes and 2.5, 2.4, and 1.7 for metabolic syndrome in the respective quartiles. Serum BDE 153 concentrations (0.04 ng/g serum) were also shown to be significantly associated with increased odds of developing gestational diabetes in 258 pregnant women from the LIFE cohort in Michigan and Texas; no associations were observed for BDE 28, BDE 47, BDE 85, BDE 99, BDE 100, or BDE 154(Smarr et al. 2016). A report from the Great Lakes Fish Consumption Study in Wisconsin found no significant difference in total serum PBDE or BDE 47 concentrations between individuals who werediagnosed with (n=64; Ln∑PBDEs reported as 0.31 ng/g) compared with non-diabetics (n=349;Ln∑PBDEs reported as 0.30 ng/g) (Turyk et al. 2015). In the Great Lakes cohort, established biomarkersof diabetes (C-reactive protein [CRP], gamma glutamyl transferase [GGT] and adiponectin) were also notassociated with PBDE exposure; however, in stratified models, Ln∑PBDE and LnBDE47 weresignificantly associated with LnGGT and Lnadiponectin in persons above the median age.In elderly populations, cross-sectional and prospective studies have found no relationship between serumPBDE (BDE 47 and BDE 153) concentrations and diabetes in cohorts from Finland (Airaksinen et al.2011) or Sweden (Lee et al. 2011). In the Finnish study, 308 participants with diabetes had median serumconcentrations of 2.7 ng/g lipid BDE 47 and 1.5 ng/g lipid BDE 153, while 1,680 nondiabetic participantshad median serum concentrations of 2.9 ng/g lipid BDE 47 and 1.7 ng/g lipid BDE 153 (Airaksinen et al. 2011). PBDE exposure levels for the Swedish study (n=725) were not available (Lee et al. 2011). PBDEs 1623. HEALTH EFFECTSIn conclusion, although numerous studies have attempted to evaluate the relationship between tissueconcentrations of PBDE and endocrine changes in humans, no clear pattern has emerged. Studies have found different results even in similar populations (e.g., adult men or pregnant women) from the same part of the world (e.g., North America or Asia) exposed to similar concentrations of PBDE. Although thespecific findings are not consistent across studies, the bulk of the data have demonstrated that PBDE caninteract with the human endocrine system to affect hormone levels.Acute-Duration Animal StudiesLower-brominated PBDEs: No exposure-related changes in thyroid weight or histology were observed in female rats exposed to penta- or tetraBDE doses up to 36 mg/kg/day for 14 days via gavage (Darnerud and Sinjari 1996; Hallgren and Darnerud 1998; Hallgren et al. 2001). No other acute studies evaluatingthyroid weight or histology were identified.Altered thyroid hormone levels have been reported following acute exposure to lower-brominated PBDEs. Significant reductions in serum T4 of 19–92% have been reported following gavage exposure to penta-, octa-, or tetraBDE at doses ≥10 and ≥0.8 mg/kg/day in rats and mice, respectively, for 1–14 days(Darnerud and Sinjari 1996; Fowles et al. 1994; Hallgren and Darnerud 1998, 2002; Hallgren et al. 2001; Hoppe and Carey 2007; Richardson et al. 2008; Stoker et al. 2004, 2005; Zhou et al. 2001). Thedecreases in serum T4 were associated with reduced ex vivo binding of T4 to the plasma thyroid hormonetransporter protein TTR (Hallgren and Darnerud 1998). Significant reductions in serum T3 were observedin rats exposed to penta- and octaBDE at doses ≥100 and ≥60 mg/kg/day, with maximum reductions up to 25–30% at 300 and 100 mg/kg/day, respectively, for 4 days via gavage (Zhou et al. 2001). No changes in serum T3 were observed in rats exposed to pentaBDE at lower doses (≤60 mg/kg/day) for 5 days viagavage (Stoker et al. 2004). No compound-related changes were observed in serum TSH levels in rats ormice exposed to penta- or octaBDE at doses up to 300 mg/kg/day for 4–14 days via gavage (Darnerudand Sinjari 1996; Hallgren and Darnerud 1998, 2002; Hallgren et al. 2001; Stoker et al. 2004; Zhou et al. 2001). In a low-dose study, male rats did not show exposure-related changes in serum T4, T3, or TSHlevels measured 45 days after a single exposure pentaBDE at doses up to 1.2 mg/kg/day via gavage (Alonso et al. 2010).Following a single gavage exposure to pentaBDE on GD 6, reductions in serum T4 levels were observedat doses ≥0.06 mg/kg in rat dams on PND 1 (23–33%), but not PND 22, and F1 males and females at0.3 mg/kg/day on PND 22 (19–23%), but not PND 1 or 14; no changes in serum T3 were observed in PBDEs 1633. HEALTH EFFECTSdams or F1 rats at any time-point at doses up to 1.2 mg/kg (Kuriyama et al. 2007). Maternal serum T4 was also reduced on GD 20 in rat dams exposed to pentaBDE doses ≥60 mg/kg/day from GD 6.5 to 19.5 via gavage, but not on GD 12 in rat dams exposed to up to 120 mg/kg/day on GDs 6.5–11.5 (Ellis-Hutchings et al. 2009). Exposure to pentaBDE at doses up to 120 mg/kg/day on GDs 6.5–11.5 or 6.5– 19.5 did not significantly alter maternal serum T3, TSH, or TTR levels (Ellis-Hutchings et al. 2009). A limited amount of information is available on hormonal effects of PBDEs other than thyroid. Therewere no clear chemical-related changes in serum corticosterone levels in female mice that were exposedto pentaBDE at doses up to 72 mg/kg/day via gavage for 14 days (Fowles et al. 1994). Limited dataregarding serum reproductive hormone levels after acute-duration exposure to lower-brominated PBDEsare discussed in Section 3.2.2.5 (Reproductive Effects).DecaBDE: No exposure-related changes in serum T4, T3, or TSH were observed in female rats exposedto decaBDE at doses up to 100 mg/kg/day via gavage for 4 days (Zhou et al. 2001). In neonatal male rats exposed to doses of 0, 6, or 20 mg/kg/day from PND 2 to 15, serum T4 was reduced by ~8 and 22% at6 and 20 mg/kg/day, respectively (Rice et al. 2007). This finding was reported as a dose-related trend;however, pair-wise statistics were not reported. No exposure-related changes in serum T4 levels were observed in similarly exposed neonatal females (Rice et al. 2007). In pregnant mice exposed to decaBDEat 0, 150, 750, 1,500, or 2,500 from GD 7 to 9 via gavage, maternal serum T4 was significantly reducedby ~11 and 14% on GD 16 in the 1,500 and 2,500 mg/kg/day groups, respectively (data reportedgraphically) (Chi et al. 2011). In the 2,500 mg/kg/day group, serum T3 levels were also significantlyreduced by ~40% (Chi et al. 2011).Intermediate-Duration Animal StudiesLower-brominated PBDEs: In a comprehensive 90-day study in rats, incidences of thyroid follicular cellhyperplasia were 0/10, 2/10, 2/10, and 5/10 in males and 0/10, 0/10, 1/10, and 4/10 in females exposed todietary pentaBDE at 0, 2, 10, and 100 mg/kg/day, respectively (WIL Research Laboratories 1984). Thethyroid hyperplasia was mild and transient, as it was characterized as very slight in severity at all doses and was no longer observed at 24 weeks postexposure in any animals. In a 28-day dietary study, thyroid hyperplasia was equivocally increased in male rats that were exposed to 90 mg/kg/day of penta- oroctaBDE (IRDC 1976). Incidences of slight or moderate hyperplasia in the 0, 9, or 90 mg/kg/day dosegroups were 0/5, 1/5, and 3/5 in pentaBDE-exposed males and 0/5, 0/5, and 3/5 in the octaBDE-exposed males, respectively; no increases were seen in females (IRDC 1976). In males rats exposed to pentaBDE PBDEs 1643. HEALTH EFFECTSat doses of 0, 3, 30, or 60 mg/kg/day for 15 days via gavage, follicular cell hypertrophy and hyperplasiawere observed in the 30 and 60 mg/kg/day groups (13/15 and 10/15, respectively); however, incidences incontrol and low-dose groups were not reported (Becker et al. 2012). Additionally, when this study wasrepeated in a different laboratory, no treatment-related histological changes were observed in the thyroidfrom the 60 mg/kg/day group (animals from 3- and 30-mg/kg/day group were not evaluated) (Becker etal. 2012). Follicular cell hypertrophy was observed in 50% of F1 rats exposed to pentaBDE at50 mg/kg/day for 18 weeks (GD 6 to PNW 16) via gavage; hypertrophy was not observed in any controls(Dunnick et al. 2012). Other histological and morphological changes observed in the thyroid included increased epithelialthickness of inner follicles of the thyroid in male rats exposed to a dietary penta-decaBDE mixture(52.1% pentaBDE, 44.2 decaBDE, and 0.4% octaBDE) at 20 mg/kg/day for 70 days (incidence data notreported) (Ernest et al. 2012); altered morphology of the epithelium (tall columnar rather than the normalcuboidal type) in 4/35 male and 1/35 female rats exposed to estimated dietary octaBDE doses of 600 and750 mg/kg/day, respectively, for 28 days (IRDC 1977); an increased incidence of cellular debris in thefollicular lumen of the thyroid in female mice exposed to dietary tetraBDE at 0.45 mg/kg/day for 28 days(Maranghi et al. 2013); a significant 60–66% increase in the follicular epithelial height score, a 23–44%decrease in the colloid area, and a 12.5–13.3% increased incidence of follicular degeneration in male andfemale rats exposed to pentaBDE at 60 mg/kg/day via gavage for 31 or 20 days, respectively (Stoker et al.2004); an increase in the observed degree of vacuolation in the thyroid of female mink exposed to 0.31 mg/kg/day for 16–17 weeks (4 weeks pre-mating through PNW 6) (Zhang et al. 2009); and aborderline significant (p=0.057) increase in thyroid follicular epithelial cell height in F1 mink exposed to 0.06 mg/kg/day via gavage during a one-generation study (4 weeks pre-mating through PNW 33) (Zhanget al. 2009). Additionally, the epithelial height of the inner follicles of the thyroids showed a biphasicresponse in rats following dietary exposure to a penta-decaBDE mixture (52.1% pentaBDE, 44.2 decaBDE, and 0.4% octaBDE) at 0. 02, 0.2, 2, or 20 mg/kg/day for 70 days (Ernest et al. 2012). Compared with controls, the epithelial height was significantly decreased at 0.02 mg/kg/day and significantly increased at 20 mg/kg/day (other doses were not significantly different from controls; quantitative data not reported) (Ernest et al. 2012). In contrast to findings in other studies, no exposure-related changes in thyroid histology were observed in male or female rats exposed to pentaBDE doses up to 200 mg/kg/day for 28 days via gavage (Van der ven et al. 2008b).Relative, but not absolute, thyroid weights were significantly elevated by 50% in male and females exposed to dietary pentaBDE at 100 mg/kg/day for 90 days (WIL Research Laboratories 1984). Eight PBDEs 1653. HEALTH EFFECTSweeks postexposure, thyroid weights were still significantly elevated by 40% in female, but not male, rats; no exposure-related changes were observed 24 weeks postexposure (WIL Research Laboratories 1984). Rats that were exposed to octaBDE in estimated dietary doses of 5, 50, or 600 mg/kg/day (males)or 7, 70, or 750 mg/kg/day (females) for 13 weeks had increased absolute and relative thyroid weights of15–31 and 32–56%, respectively, at ≥50/70 mg/kg/day (IRDC 1977). The thyroid weight increases werestill observed at 8 weeks postexposure in the 600/750 mg/kg/day groups (increased 67 and 13% in malesand females, respectively). However, no exposure-related thyroid weight changes were observed in rats exposed to pentaBDE at doses up to 200 mg/kg/day for 15–28 days via gavage (Becker et al. 2012; Vander ven et al. 2008b), in rats exposed to dietary penta- or octaBDE at doses up to 90 mg/kg/day for28 days (IRDC 1976), or in F0 or F1 mink exposed to dietary pentaBDE at doses up to 0.31 mg/kg/day in one generation studies (4 weeks pre-mating through PNW 6 or 33) (Zhang et al. 2009).Following exposure to pentaBDE doses of 0, 0.27, 0.82, 2.47, 7.4, 22.2, 66.7, or 200 mg/kg/day for28 days via gavage, hyperemia in the zona reticularis of the adrenal gland was observed in ~50% ofexposed male rats across all dose groups and occasionally in exposed female rats, but not in any controlrats (incidence data for exposed animals not reported) (Van der ven et al. 2008b). In females, there was a dose-related increase in necrotic lesions in the zona reticularis, with pyknosis at 67 mg/kg/day and widespread necrosis at 200 mg/kg/day (Van der ven et al. 2008b). The authors reported that thecombined incidence of hyperemia and zona reticularis necrosis was statistically significantly increased atthe higher doses (precise doses not specified). No histological changes were observed in the adrenalgland of female mice exposed to dietary tetraBDE at 0.45 mg/kg/day for 28 days (Maranghi et al. 2013).Mink that were exposed to dietary pentaBDE for 9 weeks had increased absolute and relative adrenalweights at 0.78 mg/kg/day (25 and 67%, respectively); no changes were observed at ≤0.63 mg/kg/day(Martin et al. 2007). No exposure-related adrenal weight changes were observed in rats exposed topentaBDE at doses up to 240 mg/kg/day for 28–31 days via gavage (Stoker et al. 2005; Van der ven et al.2008b), in rats exposed to dietary penta- or octaBDE at doses up to 90 mg/kg/day for 28 days (IRDC1976), or in male or female rats exposed to dietary octaBDE doses up to 600 and 750 mg/kg/day, respectively, for 13 weeks (IRDC 1977). Adrenal weight was not assessed in any other intermediate-duration studies identified.Serum T4 levels were significantly reduced by 22–91% in rats exposed to pentaBDE at gavage doses of≥3 mg/kg/day for 15–125 days (Becker et al. 2012; Driscol et al. 2009; Hoppe and Carey 2007; Stoker etal. 2004) or dietary pentaBDE at doses ≥20 mg/kg/day for 90 days (WIL Research Laboratories 1984). PBDEs 1663. HEALTH EFFECTSSimilarly, serum T4 levels were significantly reduced by 22% in rats exposed to a dietary penta-decaBDEmixture (52.1% pentaBDE, 44.2 decaBDE, and 0.4% octaBDE) at 20 mg/kg/day for 70 days (Ernest et al. 2012). In a 28-day study, rats were exposed to pentaBDE at 0, 0.27, 0.82, 2.47, 7.4, 22.2, 66.7, or200 mg/kg/day via gavage (Van der ven et al. 2008b). Study authors report dose-related decreases inserum T4 levels (maximal reduction of 88–89%); however, the doses at which these effects were observedwere not reported. Instead, results were reported in terms of BMD analysis (BMD/BMDLRD10%= 1.4/1.1 mg/kg/day in males and 2.7/1.8 mg/kg/day in females) (Van der ven et al. 2008b). Serum T3 levels were significantly reduced by 14–25% and serum TSH levels were significantly increased by 63– 144% in male rats exposed to pentaBDE via gavage at ≥30 mg/kg/day for 15–31 days (Becker et al. 2012;Stoker et al. 2004). However, other studies reported no significant changes in serum T3 and/or TSH levels in rats exposed to pentaBDE at doses up to 200 mg/kg/day for 20–28 days via gavage (Stoker et al. 2004;Van der ven et al. 2008b), dietary pentaBDE at doses up to 100 mg/kg/day for 90 days (WIL Research Laboratories 1984), or to a dietary penta-decaBDE mixture (52.1% pentaBDE, 44.2 decaBDE, and 0.4%octaBDE) at doses up to 20 mg/kg/day for 70 days (Ernest et al. 2012). Altered thyroid hormone levels have also been reported in F0 animals following exposure to lowerbrominated PBDEs in one-generation and gestational/lactation exposure studies. Following exposure to pentaBDE at 0, 0.5, 5, or 25 mg/kg/day via gavage for 70 days prior to mating through PND 42, serum T4 levels were significantly reduced in F0 males at ≥5 mg/kg/day (50–87%) and females at 25 mg/kg/day(67%) (Bondy et al. 2011, 2013). In rat dams exposed to pentaBDE at 18 mg/kg/day via gavage fromGD 6 to PND 18, serum T4 levels were significantly decreased by 45%; no changes were observed inserum T3, TSH, or TTR (Ellis-Hutchings et al. 2006). Following exposure to pentaBDE via gavage fromGD 6 to PND 21, maternal serum T4 levels were significantly reduced by ≥31% at ≥10.2 mg/kg/day, and maternal serum TSH levels were significantly increased by 127% at 30.6 mg/kg/day (Kodavanti et al. 2010; Zhou et al. 2002). No exposure-related changes were observed in serum T3 (Kodavanti et al. 2010;Zhou et al. 2002). In dams exposed to 30 mg/kg/day via pentaBDE-dosed cookies from GD 1 to PND 21 via dose-cookies, significant decreases were observed in serum T3 and T4 levels (quantitative data notreported); no significant changes were observed in T3 and T4 levels at ≤3 mg/kg/day or TSH levels atdoses up to 30 mg/kg/day (Bowers et al. 2015). In a similar study, a significant 25–50% decrease in maternal free and total T4 levels was observed in dams exposed to ≥10 mg/kg/day via pentaBDE-dosed cookie from GD 6 to PND 21; no exposure-related changes were observed in T3 or TSH levels at doses upto 30 mg/kg/day (Bansal et al. 2014). In rat dams exposed to tetraBDE via gavage from GD 1 to PND 14, maternal T4 levels were significantly reduced on PND 1 and 7 by 28–31% at doses ≥3.2 mg/kg/day; no changes were observed in serum T3 levels (Wang et al. 2011a). In mice, no exposure-related changes PBDEs 1673. HEALTH EFFECTSwere observed in maternal serum T4 levels after exposure to pentaBDE at 452 mg/kg/day via gavage fromGD 4 to PND 17 (Skarman et al. 2005). In one-generation dietary pentaBDE studies in mink (4 weeks premating through PNW 6), significant reductions in plasma T3, but not T4, were observed in F0 females exposed to 0.31 mg/kg/day (Zhang et al. 2009).Numerous studies have also reported decreased serum T3 and/or T4 levels in pups after gestational andlactational exposure to penta- or tetraBDE at doses as low as 0.3 mg/kg/day in rats and at 452 mg/kg/dayin mice (Bansal et al. 2014; Blanco et al. 2013; Bondy et al. 2011, 2013; Bowers et al. 2015; Ellis-Hutchings et al. 2006; Kodavanti et al. 2010; Miller et al. 2012; Poon et al. 2011; Shah et al. 2011;Skarman et al. 2005; Szabo et al. 2009; Wang et al. 2011a; Zhou et al. 2002). Changes observed in minkjuveniles included decreases in serum T3, as in the adults, but also an increase in serum T4 in juvenilefemales (Zhang et al. 2009). See Section 3.2.2.6 (Developmental Effects) for more details. Consistent, exposure-related changes were not observed between studies that evaluated serumreproductive hormone levels after intermediate-duration exposure to lower-brominated PBDEs; see Section 3.2.2.5 (Reproductive Effects) for more details. DecaBDE: Dose-related increases in thyroid hyperplasia were reported for male Sprague-Dawley rats exposed to dietary decaBDE at 80 or 800 mg/kg/day for 30 days (Norris et al. 1973, 1975a), although notin rats exposed to ≤90 mg/kg/day for 90 days, rats exposed to ≤8,000 mg/kg/day for 13 weeks, or mice exposed to ≤9,500 mg/kg/day for 13 weeks (IRDC 1976; NTP 1986). The occurrence of thyroid hyperplasia in the rats exposed to ≥80 mg/kg/day for 30 days could be related to the low puritycomposition of the older commercial decaBDE mixture tested by Norris et al. (1973, 1975a) (i.e., 77.4%decaBDE, 21.8% nonaBDE, and 0.8% octaBDE, compared to the ≥94% decaBDE composition used in the NTP studies). Similarly, observed incidences of diffuse follicular cell hypertrophy were notsignificantly elevated in rat dams exposed to dietary decaBDE from GD 10 to PND 21 at doses up to 146 mg/kg/day, compared with control (Fujimoto et al. 2011). However, in young male rats exposed to decaBDE at doses of 0, 100, 300, or 600 mg/kg/day via gavage for 33 days (PNDs 10–42), multiple areasof degenerated follicular epithelium and slight attenuation of the follicular epithelium were observed inthe thyroid glands of rat exposed to 300 or 600 mg/kg/day (incidence data not reported) (Lee et al. 2010). In young male rats exposed to decaBDE at doses of 0, 100, 300, or 600 mg/kg/day via gavage for 33 days(PND 10–42), absolute and relative thyroid weights were significantly increased by 60 and 40%, respectively, in the 600 mg/kg/day group (Lee et al. 2010). Increased absolute and relative thyroid PBDEs 1683. HEALTH EFFECTSweights were also reported in rat dams exposed to dietary decaBDE at 2, 15, or 146 mg/kg/day fromGD 10 to PND 21; findings were significant at 2 mg/kg/day (21–22% decrease) and 146 mg/kg/day (21%decrease) (Fujimoto et al. 2011). No exposure-related changes in thyroid weight were observed in rats exposed to dietary decaBDE at doses up to 90 mg/kg/day for 28 days (IRDC 1976). Unlike the lower-brominated PBDEs, serum T4 levels were not altered in rats exposed to decaBDE atgavage doses up to 600 mg/kg/day for 28–90 days (Lee et al. 2010; Van der ven et al. 2008a; Wang et al. 2010, 2011b). Serum T3 levels were significantly reduced by up to 25% in female, but not male, ratsexposed to decaBDE at 60 mg/kg/day via gavage for 28 days (Van der ven et al. 2008a), and no changesin serum T3 levels were observed in male rats exposed to decaBDE at doses up to 100 mg/kg/day viagavage for 90 days (Wang et al. 2010, 2011b). In young male rats exposed to decaBDE for 33 days(PNDs 10–45) via gavage, serum T3 was significantly reduced by up to 45% at doses ≥100 mg/kg/day and serum TSH was significantly increased by ~70% following exposure to ≥300 mg/kg/day decaBDE for33 days (PNDs 10–42) (Lee et al. 2010). In male mice, serum T4 and T3 were significantly reduced by22 and 44%, respectively, following exposure to decaBDE at gavage doses of 950 mg/kg/day for 35 days, compared with controls; no exposure-related changes in serum thyroid hormone levels were observed at750 mg/kg/day (Sarkar et al. 2015). In developing animals, significant reductions in serum T3 levels were observed following gestational and lactational exposure to decaBDE at 146 mg/kg/day in rats and 1,500 mg/kg/day in mice; no changes were observed in serum T4 or TSH (Fujimoto et al. 2011; Tseng etal. 2008); see Section 3.2.2.6 (Developmental Effects) for more details. In young male rats exposed to decaBDE at doses of 0, 100, 300, or 600 mg/kg/day via gavage for 33 days(PNDs 10–42), absolute and relative adrenal weights were significantly increased by 14 and 20%, respectively, in the 600 mg/kg/day group (Lee et al. 2010). No exposure-related changes in adrenalweight were observed in rats exposed to dietary decaBDE at doses up to 90 mg/kg/day for 28 days (IRDC1976). In mice, no significant changes in adrenal weight were observed in dams exposed to decaBDE atdoses up to 1,500 mg/kg/day from GD 0 to 17 (Tseng et al. 2008).Pancreatic effects were evaluated in rats exposed to decaBDE at doses of 0, 0.05, 1 or 20 mg/kg/day viagavage for 8 weeks (Zhang et al. 2013a). Serum insulin levels were significantly reduced by 50–60% at≥1 mg/kg/day, and blood glucose levels were elevated by 12–21% at ≥0.05 mg/kg/day (Zhang et al.2013a). Consistent with these findings, morphological changes in the pancreas were observed at≥1 mg/kg/day, including blurred boundaries among pancreatic islet cells (incidence not reported). PBDEs 1693. HEALTH EFFECTSMicroarray analysis of liver tissue indicated significant alterations in genes from the canonical pathwayfor type I diabetes mellitus (Zhang et al. 2013a). Chronic-Duration Animal StudiesLower-brominated PBDEs: No chronic-duration studies analyzing endocrine effects were located forlower-brominated PBDEs,DecaBDE: Hyperplasia of the thyroid was observed in rats and mice following repeated dietaryexposures to decaBDE. Thyroid follicular cell hyperplasia was increased in male B6C3F1 mice that were exposed to ≥94% pure commercial decaBDE for 103 weeks (NTP 1986). Incidences of the lesion were 2/50 (4%), 10/50 (20%), and 19/50 (38%) in the 0, 3,200, and 6,650 mg/kg/day dose groups of this study. Slight increases in follicular cell tumors that were considered to be equivocal evidence of thyroidcarcinogenicity were also observed in the male mice (see Section 3.2.2.7, Cancer). No decaBDE-relatedhistopathological changes in the thyroid were found after 103 weeks of exposure to ≤7,780 mg/kg/day infemale mice, ≤2,240 mg/kg/day in male Sprague-Dawley rats, or ≤2,550 mg/kg/day in female rats (NTP1986). Summary. While human data are inconsistent, they suggest that PBDEs can interact with thyroidhormone homeostasis. These data, along with available animal studies, indicate that the thyroid is a targetof concern for PBDE exposure, especially lower-brominated PBDEs. One study reported pancreatic effects, including altered insulin regulation and pancreatic lesions, following intermediate decaBDEexposure; however, no other animal studies evaluated these end points. Limited human evidence isinconclusive regarding potential associations between diabetes and PBDE exposure; however,considering the animal data, the pancreas may be a target of concern for oral PBDE exposure. There is little evidence for endocrine effects other than those mediated by the thyroid and pancreas; data foraltered reproductive hormones in humans and animals exposed to PBDEs are inconclusive (see Section 3.2.2.5, Reproductive Effects).Dermal Effects. No studies were located regarding dermal effects in humans after oral exposure toPBDEs.Histopathological examinations showed no dermal changes in rats following gavage exposure to ≤200 mg/kg/day of pentaBDE for 28 days (Van der ven et al. 2008b), dietary exposure to PBDEs 1703. HEALTH EFFECTS≤100 mg/kg/day of pentaBDE for 90 days (WIL Research Laboratories 1984), or dietary exposure to≤750 mg/kg/day of octaBDE for 13 weeks (IRDC 1977). No studies examining dermal effects afterexposure to decaBDE were located.Based on animal studies, dermal effects are unlikely with oral exposure to PBDEs.Ocular Effects. No studies were located regarding ocular effects in humans after oral exposure toPBDEs.Histopathological examinations showed no ocular effects in rats following dietary exposure to ≤100 mg/kg/day of pentaBDE for 90 days (WIL Research Laboratories 1984) or ≤750 mg/kg/day ofoctaBDE for 13 weeks (IRDC 1977). Similarly, histopathological examinations showed no ocular effectsin rats following dietary exposure to ≤1.0 mg/kg/day of 77.4% decaBDE (containing 21.8% nonaBDEand 0.8% octaBDE) for 2 years (Kociba et al. 1975; Norris et al. 1975a).Based on animal studies, ocular effects are unlikely with oral exposure to PBDEs.Body Weight Effects. The only information located was that serum concentrations of BDE 47 in agroup of 36 metabolically healthy but obese women (MHO) were not significantly different than in agroup of 40 metabolically abnormal obese women (p=33 for comparison of the medians) (Gauthier et al. 2014). The significance of this isolated finding is unknownAcute-Duration Animal StudiesLower-brominated PBDEs: Percent body weight gain was significantly decreased by 2–19% in female rats following gavage exposure to pentaBDE at 2000 mg/kg for 1 day, pentaBDE at 200 mg/kg/day for7 days, pentaBDE at 8–200 mg/kg/day for 14 days, or octaBDE at 40–200 mg/kg/day for 7 or 14 days(Bruchajzer et al. 2010, 2011, 2012). However, no changes in percent body weight gain were observed in female rats exposed to pentaBDE at doses up to 200 mg/kg/day for 7 or 14 days via gavage in anotherstudy by the same investigators (Bruchajzer 2011). Additionally, body weight effects were not reported in any other acute-duration study. No changes in body weight or body weight gain were observed in rats exposed to penta- or tetraBDE at doses up to 240 mg/kg/day for 3–14 days via gavage (Bruchajzer 2011;Hallgren and Darnerud 2002; Hallgren et al. 2001; Hoppe and Carey 2007; Stoker et al. 2004; Stoker etal. 2005; Zhou et al. 2001) or in mice exposed to penta-, tetra-, or octaBDE at doses up to 100 mg/kg/day PBDEs 1713. HEALTH EFFECTSfor 4–14 days via gavage (Hallgren and Darnerud 2002; Fowles et al. 1994; Hallgren et al. 2001; Richardson et al. 2008; Zhou et al. 2001). In pregnant rats exposed to pentaBDE at 0, 10, 100, or 200 mg/kg/day via gavage from GD 6 to 15, maternal body weight gain was significantly decreased by 20 and 30% in the 100 and 200 mg/kg/daygroups (Argus Research Laboratories 1985a). Pregnant rats exposed to octaBDE at 0, 2.5, 10, 15, 25, or50 mg/kg/day via gavage from GD 6 to 15 showed a significantly 40% reduction in maternal body weightgain at 50 mg/kg/day (WIL Research Laboratories 1986). In pregnant rabbits, exposure to octaBDE at15 mg/kg/day via gavage from GD 7 to 19 also resulted in a 7% decreased in maternal body weight gain (statistics not provided); no body weight effects were observed at ≤5 mg/kg/day (Breslin et al. 1989). In other studies, maternal body weight was not affected in rats by gestational exposure via gavage to pentaBDE at doses up to 120 mg/kg/day (Blanco et al. 2012, Ellis-Hutchings et al. 2009; Zhao et al. 2014) or tetraBDE at 0.7 mg/kg/day (Talness et al. 2008). DecaBDE: No change in body weight or body weight gains were observed in rats exposed to decaBDE atdoses up to 1,000 mg/kg/day for 4–14 days via gavage (Bruchajzer et al. 2010; Zhou et al. 2001) or in ratsand mice exposed to dietary decaBDE doses up to 16,000 and 19,000 mg/kg/day, respectively, for 7– 14 days (NTP 1986; Sakamoto et al. 2013). DecaBDE did not alter maternal body weight in miceexposed to doses up to 1500 mg/kg/day from GD 0 to 17 via gavage (Tseng et al. 2006).Intermediate-Duration Animal StudiesLower-brominated PBDEs: Percent body weight gain was significantly decreased by 8–19% in ratsexposed to pentaBDE at 8–200 mg/kg/day or octaBDE at 40–200 mg/kg/day for 21 or 28 days via gavage(Bruchajzer 2011; Bruchajzer et al. 2012). However, in another study by the same investigators thatexposed female rats to 0, 8, 40, or 200 mg/kg/day for 21 or 28 days via gavage, percent body weight gain was only significantly decreased at 200 mg/kg/day (Bruchajzer et al. 2010). Another 28-day gavagestudy reported no body weight gain changes in female rats at pentaBDE doses up to 200 mg/kg/day; however, significant changes were observed in male rats after the 28-day exposure to 0.27– 200 mg/kg/day (Van der ven et al. 2008b). The BMDRD10% and BMDLRD10% for decreased body weightgain in male rats were 61.3 and 9.7 mg/kg/day, respectively (data were reported in terms of BMD analysis only; raw data were not reported). At higher doses (≥600 mg/kg/day), rats exposed to dietary octaBDEfor 13 weeks showed ≥12% decreases in weight gain (IRDC 1977). In mink, dietary pentaBDE exposureat doses of 0.63 and 0.78 mg/kg/day for 9 weeks showed significant body weight decreases of 21 and PBDEs 1723. HEALTH EFFECTS28%, respectively; no body weight effects were observed in mink exposed to dietary concentrations of0.08 mg/kg/day (Martin et al. 2007). Body weight effects were not reported in other intermediate-duration studies. No changes in body weightor body weight gain were observed in rats exposed to pentaBDE at doses up to 250 mg/kg/day for 15– 90 days via gavage (Becker et al. 2012; Daubie et al. 2011; Hoppe and Carey 2007; IRDC 1976; Oberg etal. 2010; Stoker et al. 2004, 2005; WIL Research Laboratories 1984), in mice exposed to tetra- or diBDEat doses up to 30 mg/kg/day for 28–30 days via gavage (Wang et al. 2013; Zhang et al. 2014), in mice exposed to tetraBDE at doses up to 150 mg/kg/day for 6–12 weeks via gavage (McIntyre et al. 2015;Zhang et al. 2015a, 2015b), in rats exposed to dietary penta- or octaBDE or a penta-decaBDE mixture(52.1% pentaBDE, 44.2% decaBDE, and 0.4% octaBDE) at doses up to 90 mg/kg/day for 28–125 days(Driscol et al. 2009; Ernest et al. 2012; IRDC 1976), or in mice exposed to dietary tetraBDE at doses up to 0.45 mg/kg/day (Maranghi et al. 2013). Maternal body weight effects were not observed following gavage exposure to pentaBDE duringgestation and lactation in rats at doses up to 30.6 mg/kg/day (Bowers et al. 2015; Ellis-Hutchings et al.2006; Kodavanti et al. 2010; Zhou et al. 2002) or in mice at doses up to 452 mg/kg/day (Branchi et al. 2005; Skarman et al. 2005). In one-generation studies, no body weight effects were observed in F0 or F1rats following pentaBDE exposure to gavage doses up to ≤50 mg/kg/day (Bondy et al. 2011, 2013;Dunnick et al. 2012), in F0 dams following administration of doses up to 11.4 mg/kg/day via pentaBDEdosed vanilla wafers (Poon et al. 2011), in F0 or F1 mink exposed to dietary pentaBDE doses of≤0.31 mg/kg/day (Zhang et al. 2009), or in female F0 mice fed cornflakes dosed with 1 mg/kg/day oftetraBDE for 4 weeks prior to mating through PND 21 (Koenig et al. 2012; Ta et al. 2011)DecaBDE: One study reported significant decreases in growth in rats exposed to decaBDE via gavage for90 days (Wang et al. 2011b). “Whole-body growth rates,” defined by the study authors as the average growth rate after 90 days, were reported as 1.57, 0.59, and 0.76% in the 0, 10, and 50 mg/kg/day groups, respectively (Wang et al. 2011b). No changes in body weight or body weight gain were observed in otherstudies of rats exposed to decaBDE doses up to 1,000 mg/kg/day for 21–90 days via gavage (Bruchajzeret al. 2010; Lee et al. 2010; Van der ven et al. 2008a; Wang et al. 2010; Zhang et al. 2013a), in miceexposed to decaBDE at doses up to 1,500 mg/kg/day for 15–60 days via gavage (Heredia et al. 2012;Liang et al. 2010; Sarkar et al. 2015; Tseng et al. 2006), or in rats and mice exposed to dietary decaBDEat doses of ≤8,000 and ≤9,500 mg/kg/day, respectively, for 28–90 days (NTP 1986; Sakamoto et al. 2013;Watanabe et al. 2010a). Dietary ingestion of 77.4% decaBDE mixture (containing 21.8% nonaBDE and PBDEs 1733. HEALTH EFFECTS0.8% octaBDE) similarly caused no body weight changes in rats exposed to ≤800 mg/kg/day for 30 days or ≤1.0 mg/kg/day for 2 years (Kociba et al. 1975; Norris et al. 1973, 1975a). Mean maternal weight was suppressed by ~10% in mice exposed to dietary decaBDE doses of 290 or2,900 mg/kg/day from GD 10 to PND 21; however, the study authors did not report whether or not thisfinding was statistically significant (raw data not available for statistical analysis) (Watanabe et al.2010b). A significant 12% decrease in body weight was also reported in rat dams exposed to decaBDE at300 mg/kg/day via gavage for 3 weeks prior to mating until PND 21 (Liu et al. 2012). In other studies, decaBDE did not alter maternal body weight in rats exposed to doses up to 1,000 mg/kg/day from GD 10 to PND 21 via gavage (Biesemeir et al. 2011; Fujimoto et al. 2011) or in mice exposed to dietarydecaBDE at doses up to 260 mg/kg/day from GD 10 to PND 21 (Watanabe et al. 2008). Chronic-Duration Animal StudiesLower-brominated PBDEs: No chronic-duration studies analyzing body weight effects were located forlower-brominated PBDEs.DecaBDE: Body weight effects were not observed in rats and mice that were exposed to dietarydecaBDE doses of ≤2,550 and ≤7,780 mg/kg/day, respectively, for 103 weeks (NTP 1986). Summary. No studies are available on body weight effects of PBDEs in humans. Although some acute-and intermediate-duration animal studies reported decreased body weight, several others reported no exposure-related changes in body weight. No changes in body weight were observed in chronic studies. Based on the body of evidence from animal studies, body weight effects are unlikely to occur followingoral exposure to PBDEs at environmentally-relevant doses.Metabolic Effects. No studies were located regarding metabolic effects in humans after oral exposure to PBDEs.Animal StudiesLower-brominated PBDEs: Exposure to pentaBDE via gavage for 28 days at doses of 0, 0.27, 0.82, 2.47, 7.4, 22.2, 66.7, or 200 mg/kg/day caused significant, dose-related decreases in blood glucose levels inmale rats (Van der ven et al. 2008b). Results, reported in terms of BMD analysis, indicated a BMD10 of PBDEs 1743. HEALTH EFFECTS179.55 mg/kg/day and a BMDL10 of 66.7 mg/kg/day (raw data and statistics for individual dose groupswere not reported). No significant changes in blood glucose levels were found in female rats exposed to doses up to 200 mg/kg/day (Van der ven et al. 2008b). Reduced serum glucose levels were also reported in male rats exposed to 20 mg/kg/day of a dietary PBDE mixture containing 52.1% penta-decaBDE(DE-71), 44.2% decaBDE (BDE 209), and 0.4% octaBDE (DE-79) for 70 days; no changes wereobserved at ≤2 mg/kg/day (Ernest et al. 2012).In wild-type mice exposed to tetraBDE at 0 or 1 mg/kg/day via gavage for 6 weeks, no exposure-relatedchanges in circulating insulin levels, glucose tolerance, insulin resistance, or lipogenesis in the liver were observed (McIntyre et al. 2015). However, insulin sensitivity was significantly decreased in similarly-exposed Pten -/- mice (mice hypersensitive to insulin), compared to control Pten -/- mice. No exposure-related effects were observed in Tsc1 -/- (mice with mild insulin resistance).Exposure to pentaBDE via gavage at 250 mg/kg/day for 28 days caused hypercalcemia, hypermagnesemia, and hyperphosphatemia in male rats and hyperatremia and hypokalemia in female rats (Oberg et al. 2010). No changes in blood calcium, magnesmium, potassium, phosphorus, or sodiumlevels were observed at doses ≤25 mg/kg/day (Oberg et al. 2010). No significant changes in blood calcium, magnesium, or phosphorus levels were observed in male rats exposed to ≤20 mg/kg/day of adietary penta-decaBDE mixture containing 52.1% pentaBDE (DE-71), 44.2% decaBDE (BDE 209), and 0.4% octaBDE (DE-79) for 70 days (Ernest et al. 2012).A single study evaluated fat pad weight and adipocyte number, size, viability, lipolysis, and glucoseoxidation in male rats following exposure to pentaBDE at 14 mg/kg/day via gavage for 2 or 4 weeks(Hoppee and Carey 2007). No exposure-related effects were noted at 2 weeks. At 4 weeks, significantobservations included increased adipocyte lipolysis and decreased adipocyte glucose oxidation; however,no changes in fat pad weight or the number, size, or viability of adipocytes were observed.DecaBDE: DecaBDE exposure via gavage for 8 weeks caused elevated blood glucose levels at≥0.05 mg/kg/day in male rats (Zhang et al. 2013a). Elevated glucose levels were increased 12–21%compared with controls, and were accompanied by significantly decreased serum insulin levels at≥1 mg/kg/day; see Section 3.2.2.2, Endocrine System Effects for more details. Summary. Data regarding metabolic effects of PBDE are too limited to adequately characterize if PBDEexposure could alter metabolism in humans. PBDEs 1753. HEALTH EFFECTS3.2.2.3 Immunological and Lymphoreticular Effects Human Studies. A significant negative statistical association (not further described) was found between serum concentrations of PBDE and number of circulating lymphocytes in a subset of 18 of a cohort of33 children (18 girls and 15 boys) born in the Amsterdam/Zaandam area of the Netherlands and aged 14– 19 years at the time of this study (Leijs et al. 2009). Serum PBDE concentrations (determined as the sumof congeners 28, 47, 85, 99, 100, 153, 154, and 183) ranged from 5 to 74 ng/g lipid with a mean of13.9 ng/g lipid. Congener-specific analysis showed the main contributors to be BDE 183, BDE 154, and BDE 85. There were no other effects on leukocyte count or differential. Although dioxins and PCBswere also assessed, no association with lymphocytes was found for these chemicals. No effects onpokeweed mitogen-stimulated DNA proliferation or IgG immunoglobulin synthesis were found in human lymphocytes exposed to BDE 47 or BDE 85 in vitro (Fernlof et al. 1997). A cross-sectional study of992 subjects from Sweden aged 70 years reported that serum levels of BDE 47 (median 12.6 ng/g lipid)were significantly (p=0.03) negatively associated with blood levels of protein complement 3 (C3), but notwith serum levels of C3a or C4 (Kumar et al. 2014a); no other BDE congener was assessed. PCB levels were also negatively associated with protein complements. In a similar study of the same cohort, Kumaret al. 2014b) found no association between serum levels of BDE 47 and multiple inflammatory markers.Acute-Duration Animal StudiesLower-brominated PBDEs: Limited information is available on effects of acute-duration exposure to lower-brominated PBDEs on immunologic function in animals. A single gavage dose of 0.8–500 mg/kg pentaBDE did not affect the plaque-forming splenic cell antibody response to injected sheep red blood cells in mice (Fowles et al. 1994). Mice that were given 18, 36, or 72 mg/kg/day doses of pentaBDE via gavage for 14 days had significantly reduced antibody response to sheep red blood cells (63% of controlvalue, p<0.02) and decreased thymus weight at 72 mg/kg/day (Fowles et al. 1994). There were no exposure-related effects of the 14-day exposure to ≤72 mg/kg/day on NKC activity to murine YAC-1 target cells; NKC activity was not evaluated in the single-dose study. Another 14-day study wasconducted in which mice and rats were administered pentaBDE at 0, 18, or 36 mg/kg/day via gavage andwere evaluated for spleen and thymus weights, numbers of splenic and thymic lymphocyte subsets(CD4+, CD8+, and CD45R+ cells), and in vitro IgG immunoglobulin production in pokeweed mitogenstimulated splenocytes (Darnerud and Thuvander 1998). The only exposure-related effect in eitherspecies was significantly reduced in vitro production of IgG in pokeweed-stimulated splenocyte cultures PBDEs 1763. HEALTH EFFECTSfrom the mice exposed to 36 mg/kg/day. Mice that were similarly exposed to tetraBDE at 18 mg/kg/day via gavage for 14 days had significantly reduced numbers of total splenocytes as well as CD4+, CD8+,and CD45R+ cells in spleen (Darnerud and Thuvander 1998), but no changes in spleen or thymic weight(Hallgren et al. 2001). DecaBDE: No acute-duration studies analyzing immune effects were located for lower-brominated PBDEs.Intermediate-Duration Animal StudiesLower-brominated PBDEs: Immune function was evaluated in male mink exposed to dietary pentaBDEfor 9 weeks (Martin et al. 2007). Initial exposure concentrations were 0, 1, 10, or 100 ppm; however, after the first week of exposure, the 100 ppm group was switched to 5 ppm for weeks 2–9 because of food avoidance and weight loss concerns. Dose levels in the 1, 5/100, and 10 ppm groups were calculated to be 0.08, 0.63, and 0.78 mg/kg/day, respectively, based on estimated daily intake ranges (Martin et al. 2007). Mink were assessed weekly during exposure for antibody-mediated immunity to KLH (a carrierprotein from keyhole limpet that binds to haptens and is used to stimulate a response from the immunesystem in the form of antibody production) and at 9 weeks for PHA skin response. No dose-relatedeffects were observed for antibody production to KLH; however, at each time-point tested, antibodyproduction in the 0.63 mg/kg/day group (the 100 ppm/5 ppm group) was significantly increased compared with controls. No exposure-related changes were observed in the skin response to PHA challenge.Martin et al. (2007) also reported a significant increase in the incidence of spleen hyperplasia in minkfrom the 0.63 and 0.78 mg/kg/day groups, with 25 and 40% incidence, respectively, compared with 0%incidence in the controls. In the 0.78 mg/kg/day group, the number of germinal centers in the spleen wasalso significantly increased. In mice, histological and morphometric changes were observed in the spleenand thymus of females exposed to dietary tetraBDE at 0.45 mg/kg/day for 28 days (only dose tested;males not evaluated) (Maranghi et al. 2013). In the spleen, follicular hyperplasia with germinal centerdevelopment was observed in 9/10 exposed mice, compared with 4/10 controls, and lymphocyticinfiltration involving the red pulp was observed in 5/10 exposed mice, compared with 0/10 controls. In the thymus, 7/9 exposed mice showed Hassal’s bodies, compared with 2/10 controls, and 5/9 showed lymphocytic apoptosis, compared with 0/10 controls. The ratio between area of cortex and medulla in thethymus was significantly increased by 43% in exposed mice, compared with controls. In the thyroid, cellular debris was observed in the follicular lumen of 5/7 exposed mice, compared with 0/10 controls PBDEs 1773. HEALTH EFFECTS(Maranghi et al. 2013). In other studies, no exposure-related changes were observed in spleen, thymus,lymph node, and/or bone marrow tissue histology in rats exposed to pentaBDE at gavage doses up to 250 mg/kg/day for 28 days (Oberg et al. 2010; Van der ven et al. 2008b), pentaBDE at dietary doses up to 100 mg/kg/day for 90 days (WIL Research Laboratories 1984), or octaBDE at dietary doses up to750 mg/kg/day for 13 weeks (IRDC 1977). Additionally, no exposure-related changes were observed inthe histology of the spleen, thymus, Peyer’s patches, or mesenteric lymph nodes in rat dams exposed topentaBDE at doses up to 25 mg/kg/day via gavage for 70 days prior to mating through PND 21 (~21 weeks) (Bondy et al. 2013). In offspring, a significant dose-related trend was observed in theincidence of apoptotic lymphocytes and tingible macrophages in the thymus of PND 43 males, but notfemales (Bondy et al. 2013); see Section 3.2.2.6 (Developmental Effects) for more details. Martin et al. (2007) reported a significant 29% increase in relative, but not absolute, spleen weight inmink exposed to dietary pentaBDE at 0.78 mg/kg/day for 9 weeks. In a 28-day study, rats were exposed to pentaBDE at 0, 0.27, 0.82, 2.47, 7.4, 22.2, 66.7, or 200 mg/kg/day via gavage (Van der ven et al. 2008b). Study authors reported a dose-related decrease in thymus weight in male rats (maximal decrease of 19.4%); however, the lowest dose at which the effect was observed was not reported. Instead, resultswere reported in terms of BMD analysis (BMD/BMDLRD20%=194.2/110 mg/kg/day). No dose-relatedchanges were reported for female thymus weight, male or female spleen weight, or T-cell, B-cell, ormacrophage population distribution in the spleen in either sex (Van der ven et al. 2008b). In otherstudies, no exposure-related changes were observed in spleen or thymus weights in rats exposed to pentaor diBDE at gavage doses up to 250 mg/kg/day for 28 days (Oberg et al. 2010; Zhang et al. 2014), in ratsexposed to pentaBDE at gavage doses up to 0.015 mg/kg/day for 90 days (Daubie et al. 2011), in ratsexposed to dietary penta- or octaBDE at doses up to 100 mg/kg/day for 28 or 90 days (IRDC 1976; WILResearch Laboratories 1984), in rats exposed to a dietary penta-decaBDE mixture (52.1% pentaBDE[DE-71], 44.2% decaBDE [BDE 209], 0.4% octaBDE [DE-79]) at doses up to 20 mg/kg/day for 70 days(Ernest et al. 2012), in rats exposed to dietary octaBDE at doses up to 750 mg/kg/day for 13 weeks (IRDC1977), or in mice exposed to dietary tetraBDE at 0.45 mg/kg/day for 28 days (Maranghi et al. 2013). Additionally, no exposure-related changes in spleen or thymus weights were observed in rat dams oroffspring exposed to pentaBDE at 18 mg/kg/day via gavage from GD 6 to PND 21 (Ellis-Hutchings et al. 2006), in rat dams exposed to pentaBDE at doses up to 25 mg/kg/day via gavage for 70 days prior to mating through PND 21 (~21 weeks) (Bondy et al. 2013), or in mink sows exposed to dietary pentaBDEat doses up to 0.31 mg/kg/day for 4 weeks prior to mating through PNW 6 (Zhang et al. 2009). Although immune function was not altered, increased thymus weights, as well as altered serum immunoglobin levels and lymphocyte proliferation, were observed in the offspring of rat dams exposed to pentaBDE at PBDEs 1783. HEALTH EFFECTSdoses ≥5 mg/kg/day via gavage for 70 days prior to mating through PND 21 (~21 weeks) (Bondy et al.2013); see Section 3.2.2.6 (Developmental Effects) for more details. In a poorly-reported study, immune end points were evaluated in mouse dams and offspring followingexposure to pentaBDE at 0, 50, 100, or 200 mg/kg/day via gavage from GD 6 to PND 21 (Hong et al. 2010). The absolute and relative spleen weights were significantly decreased in the 100 and 200 mg/kg/day groups and the thymi of exposed dams were “weighed decreasingly” (quantitative data were not reported). No histopathological alterations were observed in the spleen or thymus. The studyabstract reports decreased cellularity of the spleen and thymus in dams; however, cellularity data for the dams are not reported in the results section of the paper. No statistically significant increases in T- orB-cell lymphocyte proliferation were observed, and no exposure-related changes in T-cell, B-cell, ormacrophage population distribution were observed in the spleen. The study authors reported decreases inserum IgM levels, but statistical significance was not reported. No changes in serum IgG1 were observed. The study authors also noted some immune effects in offspring, including decreased spleen weight and cellularity and altered serum immunoglobin levels on PND 21 in groups exposed to ≥100 mg/kg/day and increased T-cell proliferation following in vitro exposure to concanavalin A (ConA)at 200 mg/kg/day on PND 63; see Section 3.2.2.6 (Developmental Effects) for more details. DecaBDE: One week following exposure to dietary decaBDE at 0 or 1,800 mg/kg/day for 28 days, micewere intranasally infected with the RSV (Watanabe et al. 2010a). No exposure-related differences inpulmonary viral titers were observed 5 days post-infection (Watanabe et al. 2010a). In contrast, pulmonary viral loads were increased post-infection in PND 28 offspring of mouse dams exposed to dietary decaBDE at doses ≥260 mg/kg/day from GD 10 to PND 21 (Watanabe et al. 2008, 2010b); see Section 3.2.2.6 (Developmental Effects) for more details. In another high-dose study, female mice exposed to decaBDE at 800 mg/kg every other day showed impaired CD4 T-cell function from 4 to 10 months of exposure, compared with controls (Feng et al. 2016b). Significant alterations in peripheral CD4 T-cells from exposed mice included significant decreases in in vitro cytokine production of TNF-α,IFN-γ, and IL-2, decreased percentage of multi-functional CD4 T-cells (cells capable of producingmultiple cytokines following mitogen production), increased percentage of T regulatory cells(CD4+CD25+Foxp3+), decreased percentage of proliferating CD4 T-cells, and decreased percentage ofantigen-specific CD4 T cells following in vivo Listeria monocytogenes infection. In rats exposed to decaBDE at 0 or 300 mg/kg/day via gavage for 21 days prior to mating through PND 21, the distribution of T-lymphocytes in the thymus was significantly altered in exposed rat dams, PBDEs 1793. HEALTH EFFECTSwith significantly decreased CD3+, CD4+, CD4+/CD8+, CC3+/CD8+, and CD3+/CD4+ T-cells andsignificantly increased CD4-/CD8- T-cells compared with control dams (Liu et al. 2012). CD161+ NKCs were also significantly decreased, compared with controls. Additionally, the lymphocyte stimulation index in response to in vitro PHA exposure was significantly reduced by ~2-fold in lymphocytesharvested from exposed rat dams, compared with control dams (Liu et al. 2012). In another study, no dose-related changes were reported for T-cell, B-cell, or macrophage population distribution in the spleen of rats exposed to decaBDE at doses up to 60 mg/kg/day via gavage for 28 days (Van der ven et al. 2008a). Following exposure to decaBDE at 300 mg/kg/day via gavage for 21 days prior to mating through PND 21, serum IgM and IgG were significantly decreased by 28 and 4%, respectively, compared with controls (Liu et al. 2012). No exposure-related changes were observed in IFN- γ or IL-4 (Liu et al. 2012). Histopathological examinations of spleen, thymus, lymph node, and/or bone marrow tissues showed no effects in rats or mice exposed to dietary decaBDE at doses up 8,000 or 9,500 mg/kg/day, respectively, for 13 weeks (NTP 1986). In rat dams exposed to decaBDE at 0 or 300 mg/kg/day via gavage for 21 daysprior to mating through PND 21 (~11 weeks), the exposed group showed significantly increased “histological scores” in the spleen and thymus compared with controls (methods of histological scoringand incidences of lesions were not reported) (Liu et al. 2012). Lesions observed in the thymus of exposed rats included thickened thymus capsule, decreased lymphoid tissue in the cortex with adipose tissue replacement, increased medulla size, and obscured corticomedullary junction. The spleen showed decreased size and number of lymphoid nodules, thinner lymphatic sheath around arteries, and fibrotictissue with macrophages in the medulla.In a 28-day study, rats were exposed to decaBDE at 0, 1.87, 3.45, 7.5, 15, 30, or 60 mg/kg/day via gavage(Van der ven et al. 2008a). The study authors reported a dose-related decreased in thymus weight infemale rats (maximal decrease of 16.6%); however, the lowest dose at which the effect was observed wasnot reported. Instead, results were reported in terms of BMD analysis (BMD/BMDLRD20%=75/43 mg/kg/day). No dose-related changes were reported for male thymus weight or male or female spleen weight (Van der ven et al. 2008a). Absolute and relative spleen weights were significantlydecreased by 12 and 27%, respectively, in rat dams exposed to decaBDE at 300 mg/kg/day via gavage for21 days prior to mating through PND 21 (~11 weeks); no exposure-related changes were observed inthymus weight (Liu et al. 2012). In other studies, no exposure-related changes in spleen weight were observed in rats exposed to decaBDE at doses up to 20 mg/kg/day via gavage for 8 weeks (Zhang et al. PBDEs 1803. HEALTH EFFECTS2013a) or in mouse dams or PND 71 male offspring exposed to decaBDE at doses up to 1,500 mg/kg/dayfrom GD 0 to 17 via gavage (Tseng et al. 2008).In a poorly-reported study, immune end points were evaluated in mouse dams and offspring followingexposure to decaBDE at 0, 500, 2,500, or 12,500 mg/kg/day via gavage on GD 6 to PND 21 (Hong et al. 2010). No exposure-related changes were reported for maternal spleen or thymus weight or histology.T-cell lymphocyte proliferation (in response to ConA) was “slightly” increased. No statisticallysignificant increases in B-cell lymphocyte proliferation were observed and no exposure-related changes inT-cell, B-cell, or macrophage population distribution were observed in the spleen. The study authorsreported increased serum IgG1 and IgM levels; however, statistics were not reported and changes do notappear dose-related in graphically-presented data. In offspring, altered spleen cell populations were noted on PND 21; see Section 3.2.2.6 (Developmental Effects) for more details. Chronic-Duration Animal StudiesLower-brominated PBDEs: No chronic-duration studies analyzing immune effects were located forlower-brominated PBDEs.DecaBDE: Following chronic ingestion of decaBDE for 103 weeks, an increased incidence of splenichematopoiesis was observed in female rats at ≥1,200 mg/kg/day (12/49, 24/48, and 17/50 at 0, 1,200, and 2,550 mg/kg/day, respectively); only the incidence in the mid-dose group was statistically significant. Inmale rats, the incidence of splenic fibrosis was significantly increased at 2,240 mg/kg/day (5/49, 8/50, and 13/49 at 0, 1,120, and 2,240 mg/kg/day, respectively) (NTP 1983). The incidence of lymphoid hyperplasia in the mandibular lymph node was also significantly increased in male rats at2,240 mg/kg/day (4/50, 6/50, and 13/49 at 0, 1,120, and 2,240 mg/kg/day, respectively) (NTP 1983). No exposure-related histopathological lesions were observed in male or female mice following chronic ingestion of decaBDE for 103 weeks at doses up to 6,650 and 7,780 mg/kg/day, respectively (NTP 1983).Summary. Evidence from animals suggest that PBDE exposure may cause immune suppression,particularly in infants or children (see Section 3.2.2.6, Developmental Effects for more details), but dataare limited and inconsistent. Additionally, comprehensive immunological evaluations have not been performed and human data are limited. Therefore, currently available information is insufficient toadequately characterize the human immunotoxic potential of PBDEs. The highest NOAEL values and all PBDEs 1813. HEALTH EFFECTSLOAEL values from each reliable study for immunological effects in each species and duration categoryare recorded in Table 3-2 and plotted in Figure 3-2.3.2.2.4 Neurological Effects Human Studies. No association was found between serum PBDE concentrations (BDE 28, BDE 47, BDE 66, BDE 85, BDE 99, BDE 100, BDE 138, BDE 153, BDE 154, and their sum) and neuropsychological function assessed by 34 tests of cognitive and motor function, affective state, and olfactory function in a study population of 144 volunteers (67 males and 77 females) between the ages of55 and 74 who lived for at least 25 years in the upper Hudson valley of New York State (Fitzgerald et al. 2012). Median total serum PBDE concentration in the study population was 23.9 ppb (lipid weight), with the detection limit exceeded in 89% of the population. Neurobehavioral function in adolescents wasstudied by Kicinski et al. (2012). The study population included 515 secondary school students(271 boys) from Flanders, Belgium, with a mean age of 14.9 years. The students were given a computerized battery of neurological tests. PBDE concentrations (BDE 28, BDE 47, BDE 99, BDE 100, BDE 153, BDE 154, BDE 183, and BDE 209) were measured in the serum. The median concentration oftotal PBDE in the serum was 7 ng/L. Serum PBDE concentrations were not associated with most aspects of neurological performance investigated, but there was a significant deterioration in performance in thefinger tapping test with increasing PBDE level, suggesting an effect of PBDE on motor activity. Studiesof neurological function in relation to PBDE concentrations in young children are discussed in Section3.2.2.6 on Developmental Effects.Acute-Duration Animal StudiesLower-brominated PBDEs: In a neurotoxicity screen, adult male rats were exposed once to pentaBDE at0, 0.6, or 1.2 mg/kg via gavage (Belles et al. 2010). Rats were assessed using a functional observation battery 3, 21, and 44 days after pentaBDE administration. After 45 days, rats were assessed in a varietyof behavioral tests over 9 days, including open-field activity, passive avoidance test, and the Morris watermaze. Following completion of behavioral tests, rats were sacrificed and brains were removed forbiochemical analysis of oxidative stress markers (right hemisphere) and histopathological (left hemisphere) examinations of the cortex, hippocampus, and cerebellum. No exposure-relatedneurobehavioral, histological, or biochemical effects were observed. PBDEs 1823. HEALTH EFFECTSIn developing animals, a series of studies evaluated neurobehavior at 2–8 months of age following single-day exposure to lower-brominated PBDEs on PND 3, 10, or 19 in rats and mice. Collectively, thesestudies showed altered open-field activity, impaired habituation and altered learning and memory in ratsand mice exposed to penta-, hexa-, tetra-, or octaBDE at doses as low as 0.4, 0.45, 1, or 16.8 mg/kg, respectively (Eriksson et al. 2001, 2002b, 2006; Fischer et al. 2008; Hallgren et al. 2015; He et al. 2009, 2011; Gee and Moser 2008; Sand et al. 2004; Viberg et al. 2002, 2003a, 2004a, 2004b, 2005, 2006). No changes in open-field behavior were observed in mice exposed to heptaBDE at 15.2 mg/kg or nonaBDEat 18.5 on PND 10 (Viberg et al. 2006). Besides behavioral changes, other significant neurodevelopmental effects observed in these studies included ultrastructural changes, altered nicotinic receptordensity, and altered gene and protein expression levels in the hippocampus. Altered open-field behaviorwas also observed in PND 36 offspring following acute gestational exposure to pentaBDE at doses as lowas 0.06 mg/kg via gavage on GD 6 (Kuriyama et al. 2004). See Section 3.2.2.6 (Developmental Effects)for more details regarding neurodevelopmental effects of acute exposure to lower-brominated PBDEs. DecaBDE: No studies evaluating neurobehavior or neuropathology in adult rats or mice following acuteexposure to decaBDE were identified; however, there were no overt signs of neurotoxicity in rats andmice exposed to decaBDE in estimated dietary doses of ≤16,000 and ≤19,000 mg/kg/day, respectively, for 14 days (NTP 1986). In developing animals, a series of studies reported decreased spontaneous activity and impairedhabituation in open-field testing in 2–6-month-old mice that were exposed to decaBDE on PND 3 atdoses as low as 2.22 mg/kg (Buratovic et al. 2014; Johansson et al. 2008; Viberg et al. 2003b, 2007). These changes in open-field behaviors were not observed in animals exposed on PND 10 or 19. In contrast, significantly increased locomotor activity during the first 1.5 hours of a 2-hour observation period was observed in PND 70 males following exposure to decaBDE at 20 mg/kg/day via micropipettefrom PND 2 to 15 (Rice et al. 2007). No changes in locomotor activity were observed at PND 70 in females or at 1 year in either sex at doses up to 20 mg/kg/day (Rice et al. 2007). In operant training and visual discrimination tasks, learning impairments and impulsivity were observed in aging mice(16 months old), but not young adult mice (3 months old), that had been exposed to decaBDE at20 mg/kg/day from PND 2 to 15 (Rice et al. 2009). See Section 3.2.2.6 (Developmental Effects) for moredetails regarding neurodevelopmental effects of acute exposure to decaBDE. PBDEs 1833. HEALTH EFFECTSIntermediate-Duration Animal StudiesLower-brominated PBDEs: No exposure-related changes were observed in open-field behavior, anxiety-like behavior in the elevated plus-maze performance, or learning and memory in the Morris water maze inmale rats following exposure to pentaBDE at doses up to 0.015 mg/kg/day (highest dose tested) viagavage for 90 days (Daubie et al. 2011). In a higher dose study, impaired attention and inhibitory controlwas observed in a series of 5-choice serial reaction time tasks assessed from PND 40 to 125 in male mice exposed to dietary pentaBDE at 26.2 mg/kg/day from PND 1 to 125; no exposure-related changes were observed at 17.5 mg/kg/day (Driscoll et al. 2009). In adult male rats exposed to tetraBDE at 0, 0.1, 0.5, or1 mg/kg/day via gavage for 30 days, rats in all exposure groups showed impaired learning and memory in the Morris water maze (Yan et al. 2012). Exposed rats required significantly more time to find the hidden platform in the Morris water maze compared with control group, without showing any differences inswim speed. Additionally, significant decreases in the time spent in the target quadrant and the number ofcrossings over the original platform location were observed during the retention trial on day 5 (Yan et al. 2012).No exposure-related changes in brain weight and/or histology were observed in rats exposed to pentaBDEat gavage doses up to 250 mg/kg/day for 28 days (Oberg et al. 2010; Van der ven et al. 2008b), in ratsexposed to pentaBDE at gavage doses up to 0.015 mg/kg/day for 90 days (Daubie et al. 2011), in ratsexposed to penta- or octaBDE at dietary doses up to 750 mg/kg/day for 28 or 90 days (IRDC 1976, 1977;WIL Research Laboratories 1984), in rat dams or pups exposed to pentaBDE at a gavage dose of18 mg/kg/day from GD 6 to PND 18 (Ellis-Hutchings et al. 2006), or in mink sows or kits exposed to pentaBDE at dietary doses up to 0.31 mg/kg/day for 4 weeks prior to mating through PNW 6 or 33 (Zhang et al. 2009).The density of NMDA receptor subunits, NR1 and NR2B, and the glutamate receptor, Glu, wasdetermined in the CA1, CA3, and dentate gyrus of the hippocampus in adult male rats exposed to tetraBDE at 0, 0.1, 0.5, or 1 mg/kg/day via gavage for 30 days (Yan et al. 2012). Immunohistochemicalstaining showed significant decreases in the density of NR1 and Glu in the hippocampus of all exposed rats and NR2B in the hippocampus of rats exposed to 0.5 and 1 mg/kg/day. Additionally, significantdecreases in hippocampal mRNA levels were observed for NR1 and NR2C in all dose groups and NR2D in the 0.5 and 1 mg/kg/day groups. No exposure-related changes were observed in NR2A or NR2B mRNA levels (Yan et al. 2012). PBDEs 1843. HEALTH EFFECTSCholinergic effects were evaluated in mink sows and kits exposed to dietary pentaBDE at 0, 0.01, 0.05, or0.25 mg/kg/day from pre-mating day 28 to PNW 6 (Bull et al. 2007). No exposure-related effects were observed in acetylcholinesterase concentration (ACh), acetylcholinesterase activity (ChE), or muscarinic receptor (mAChR) or nicotinic receptor (nAChR) binding in the cerebral cortices of sows or kits. Plasma ChE activity in sows from the 0.25 mg/kg/day group was significantly increased by 3-fold, compared with controls. Plasma ChE activity was not significantly correlated with cortical ChE activity; however, it was significantly correlated with absolute and relative liver weights. Therefore, altered ChE activity inplasma in the high-dose sows may reflect exposure-related effects in the liver, rather than the centralnervous system. No changes were observed in plasma ChE activity in kits (Bull et al. 2007).Several studies have reported delayed ontogeny of reflexes, neurobehavioral impairments, and ultrastructural and biochemical changes in the hippocampus of offspring after gestational and lactationalexposure to penta- or tetraBDE in rats and mice at doses as low as 2 and 0.03 mg/kg/day, respectively(Blanco et al. 2013; Bowers et al. 2015; Branchi et al. 2001, 2002, 2005; Cheng et al. 2009; Koenig et al. 2012; Ta et al. 2011; Woods et al. 2012). See Section 3.2.2.6 (Developmental Effects) for more detailsregarding neurodevelopmental effects of intermediate-duration exposure to lower-brominated PBDEs.DecaBDE: Following a 15-day exposure to decaBDE at 20 mg/kg/day via gavage, male mice showeddecreased anxiety behavior in the elevated zero maze, including decreased latency to first entry into the open region, increased time spent and number of entries into the open region, and increased number ofhead dips, compared with controls (Heredia et al. 2012). No exposure-related changes were observed inanxiety behaviors in the light/dark test, in learning or memory in the Morris water maze, or in generalneurological behaviors assessed using a functional observation battery (Heredia et al. 2012). In anotherstudy, no changes were observed in open-field behavior of male rats exposed to decaBDE at doses up to 50 mg/kg/day via gavage for 90 days (Wang et al. 2011b). No overt signs of neurotoxicity were observed in rats and mice exposed to decaBDE in estimated dietary doses of ≤8,000 and ≤9,500 mg/kg/day, respectively, for 13 weeks (NTP 1986). Rats exposed to decaBDE at dietary doses up to 90 mg/kg/day for 28 days showed no change in brain weight (IRDC 1976). No exposure-related changes in brain weight or AchE activity were observed inmice exposed to decaBDE at doses up to 160 mg/kg/day via gavage for 15, 30, or 60 days (Liang et al. 2010). PBDEs 1853. HEALTH EFFECTSThere is limited evidence for neurodevelopmental effects following gestational and lactational exposure to decaBDE. No neurobehavioral changes were observed in rat offspring exposed to decaBDE at dosesup to 1,000 mg/kg/day from GD 6 to PND 21 (Biesemeier et al. 2011). However, impaired spatiallearning was observed in offspring of rat dams exposed to decaBDE at doses ≥30 mg/kg/day from GD 1to 14 (Chen et al. 2014). Additionally, altered hippocampal electrophysiology was observed in mice offspring exposed to 20.1 mg/kg/day from GD 1 to PND 41 (Xing et al. 2009) and altered hippocampalimmunohistochemistry was observed in mice offspring exposed to ≥15 mg/kg/day from GD 10 toPND 21 (Fujimoto et al. 2011). See Section 3.2.2.6 (Developmental Effects) for more details regardingneurodevelopmental effects of intermediate-duration exposure to decaBDE.Chronic-Duration Animal StudiesLower-brominated PBDEs: No chronic-duration studies analyzing neurological effects were located forlower-brominated PBDEs.DecaBDE: There were no indications of neurotoxicity in rats and mice in lifetime feeding studies ofdecaBDE at doses as high as 2,550 and 7,780 mg/kg/day, respectively, as assessed by overt clinical signs and nervous system histopathology (NTP 1986). Although the high doses and extended exposuredurations in the NTP (1986) studies provided opportunities for the induction and/or development ofeffects, neurotoxicity is incompletely evaluated due to the lack of testing for subtle behavioral and othersensitive neurological end points. Summary. While the nervous system is a target of concern during early development (see Section 3.2.2.6, Development Effects for more details), it is unclear if the developed nervous system is a target of oral PBDE toxicity. Animal data suggest the oral PBDE exposure may lead to neurobehavioralchanges; however, available information is insufficient to adequately characterize the neurotoxic potentialof PBDEs in adults and adolescents. The highest NOAEL values and all LOAEL values from eachreliable study for neurological effects in each species and duration category are recorded in Table 3-2 and plotted in Figure 3-2.3.2.2.5 Reproductive EffectsHuman Studies. Two studies have demonstrated reproductive effects in men associated with exposure toPBDEs. In a small study limited to 10 young adult Japanese males, strong, statistically significant inverse PBDEs 1863. HEALTH EFFECTScorrelations were found between serum levels of BDE 153 and sperm concentration and testis size(Akutsu et al. 2008). Both blood and sperm were collected monthly and pooled for each participant over1 year. BDE 153 concentrations for the 10 participants ranged from 0.37 to 1.1 ng/g lipid. There was no relationship to sperm concentration or testes size for other PBDE congeners or for the sum of the mostprevalent congeners, which included BDE 153 (BDE 47, BDE 99, BDE 100, and BDE 153) and ranged from 1.1 to 8.6 ng/g lipid for the 10 participants. Among a group of 52 Canadian men recruited at afertility clinic, sperm mobility was significantly reduced in association with increased serum PBDEconcentrations (BDE 47, BDE 100, and total) (Abdelouahab et al. 2011). Results for BDE 153 weresimilar, but not statistically significant. Relationships between PBDE and sperm concentration were consistently negative as well, but also not significant. The median ∑PBDE concentration in thispopulation was 0.302 ng/mL and the maximum was 2.250 ng/mL. A study of 468 male partners ofcouples discontinuing contraception for purposes of becoming pregnant from Michigan and Texas foundthat, in general, serum concentrations PBDEs were negatively associated with parameters of semenquality, specifically increased percentages of abnormal morphology (Mumford et al. 2015). The studyevaluated 10 BDEs: BDE 17, BDE 28, BDE 47, BDE 66, BDE 85, BDE 99, BDE 100, BDE 153, BDE 154, and BDE 183. BDE 153, however, was associated with increased sperm concentration. The95th percentile serum concentrations (unadjusted for lipid content) for the BDEs analyzed ranged from0.007 ng/g serum for BDE 183 to 0.761 ng/g serum for BDE 47. Significant associations at the α=0.01 level were found only for BDE 17 (increased sperm with coiled tail) and for BDE 28 (increasedbicephalic sperm). A cross-sectional study that included 99 men from Greenland, 100 from Poland, and 100 from Ukraine found no association between serum levels of BDE 47 and BDE 153 with parametersof semen quality including markers of DNA damage and apoptosis (Toft et al. 2014). Among the threegroups, median concentrations of BDE 47 ranged from 0.2 to 2.0 ng/g lipid and median concentrations ofBDE 153 ranged from 0.3 to 2.7 ng/g lipid. Studies of hormone levels in men in relation to PBDE concentrations have not produced consistentresults. Concentrations of PBDE (BDE 47, BDE 99, and BDE 100) in house dust samples weresignificantly associated with changes in some sex hormone levels (decreased FSH, LH, and free androgenindex, and increased Inhibin B and sex hormone binding globulin [SHBG]), although not testosterone orestradiol, in 24 adult men recruited through a Massachusetts infertility clinic (Meeker et al. 2009).Median and maximum concentrations of PBDE in dust collected from houses of study subjects were 500 and 7,620 ng/g dust for BDE 47, 838 and 9,220 ng/g dust for BDE 99, and 180 and 2,830 ng/g dustfor BDE 100. A larger subsequent study by the same researchers with 38 additional subjects using the same design found a significant negative correlation with FSH and significant positive correlations with PBDEs 1873. HEALTH EFFECTSestradiol and SHBG for pentaBDE (sum of BDE 47, BDE 99, and BDE 100; median and maximumconcentrations of 1,049 and 22,300 ng/g in dust), significant positive correlations with testosterone andLH for octaBDE (sum of BDE 183 and BDE 201; median and maximum concentrations of 30.5 and 1,181 ng/g in dust), and a significant negative correlation with testosterone for decaBDE (sum of BDE206, BDE 207, BDE 208, and BDE 209; median and maximum concentrations of 1,800 and 38,483 ng/gin dust) (Johnson et al. 2013). No relationships were found between serum concentrations of BDE 47 and testosterone, FSH, LH, or prolactin in adult male Baltic Sea area residents of Sweden and Latvia with a wide range of fish consumption behavior and BDE 47 concentrations ranging from 0.10 ng/g lipid at the10th percentile to 5.16 ng/g lipid at the 90th percentile (median=1.04 ng/g lipid) (Hagmar et al. 2001). In astudy of serum PBDE and hormone levels in adult male sport fish consumers from the Great Lakes, a significant positive relationship was found for BDE 47 (but not other congeners) and serum testosterone levels (Turyk et al. 2008). BDE 47 concentrations in this population ranged from 0.01 to 5.90 ng/g(median=0.11 ng/g), while ∑PBDE concentrations ranged from 0.13 to 10.15 ng/g (median=0.26 ng/g).Serum concentrations of FSH, LH, or estradiol were not associated with serum levels of BDE 47 orBDE 153 in a study of 299 spouses of pregnant women from Greenland, Poland, and Ukraine (Toft et al. 2014). Median concentrations of BDE 47 ranged from 0.2 to 2.0 ng/g lipid and median concentrations ofBDE 153 ranged from 0.3 to 2.7 ng/g lipid. In women, Chao et al. (2007) found no significant relationship between concentrations of PBDE in breastmilk and pre-pregnancy menstrual cycle length (by questionnaire) in an analysis of 20 pregnant women from Taiwan with median PBDE concentrations in breast milk of 3.65 ng/g lipid, predominantly BDE 47 and BDE 153. However, in a larger study of 46 pregnant Taiwanese women recruited several years later,Chao et al. (2010) found significant increases in average length of pre-pregnancy menstrual cycle withincreased breast milk concentrations of PBDE (total and multiple individual congeners). Median breastmilk total PBDE concentration in this study group was 2.84 ng/g lipid, with predominant congeners beingBDE 47, BDE 153, and BDE 209. A study of 42 Cree women of James Bay, Canada, reported thatconcentrations of BDE 47 and BDE 153 in plasma were associated with increased menstrual cycle length(Wainman et al. 2016). Geometric mean plasma concentrations of BDE 47 and BDE 153 were 14.1 and 4.5 ng/g lipid, respectively, which according to the investigators, were comparable to those reported forwomen 20–39 years old from the general Canadian population; serum cadmium and selenium were alsoassociated with increased cycle length. In the study by Chao et al. (2010), age at menarche was notrelated to breast milk PBDE concentrations, but an analysis of data from a sample of 271 adolescent U.S.girls age 12–19 years with serum total PBDE concentrations ranging from 6.4 to 636.5 ng/g lipid (median=44.7 ng/g lipid) from NHANES (2003–2004) found that higher serum PBDE concentrations PBDEs 1883. HEALTH EFFECTSwere associated with younger age of menarche in this population (Chen et al. 2011). There was no relationship between mid-pregnancy serum PBDE concentrations and pre-pregnancy menstrual cycling ina cohort of 223 pregnant low-income women from the Salinas Valley of California (part of theCHAMACOS study) (Harley et al. 2010), but there were significant decreases in fecundability(i.e., increases in time to pregnancy between stopping contraception and becoming pregnant) associated with increased serum concentrations of BDE 47 (geometric mean=14.9 ng/g lipid), BDE 99 (geometicmean=4.4 ng/g lipid), BDE 100 (geometic mean=2.8 ng/g lipid), BDE 153 (geometic mean=2.5 ng/glipid), and their sum in this population and/or a subset comprising 107 women actively trying to becomepregnant. However, fecundability was not significantly related to serum PBDE concentrations(concentrations not reported) in a cohort of 501 Michigan and Texas couples followed prospectively for1 year after discontinuing contraception for the purpose of becoming pregnant (Buck Louis et al. 2013). Blood was collected from both male and female partners in this study and fecundity of the couple wasassessed in relation to each partner’s exposure. A study of 65 women undergoing in vitro fertilizationfound no association between serum PBDE concentrations (median=12.6 ng/g lipid, maximum=113 ng/glipid) and risk of implantation failure, but did find a significantly increased risk of failure associated withincreased (i.e., detectable) concentrations of BDE 153 (but not other congeners or total PBDE) in follicular fluid (Johnson et al. 2012). No significant relationship was found between concentrations of thePBDE metabolite 6-OH-BDE-47, which ranged from <4 to 127 pg/g wet weight, with a median of26 pg/g wet weight) and 17β-estradiol in umbilical cord serum collected from 26 births in South Korea(Wan et al. 2010). Serum PBDE concentrations had no effect on expression of genes for the sex steroid enzymes aromatase (CYP19A1) and 17-α-hydroxylase or estrogen receptors α and β (ESR1 and ESR2) in leukocytes collected from 139 adult daughters of Michigan fisheaters with serum total PBDEconcentrations ranging from 4.3 ng/g lipid at the 5th percentile to 209.5 ng/g lipid at the 95th percentile (median=33.8 ng/g lipid) (Karmaus et al. 2011).Vagi et al. (2014 examined the association between blood levels of multiple environmental pollutants and Polycystic Ovary Syndrome (PCOS). Fifty-two PCOs patients and 50 controls were included in thestudy. PCOS is an endocrine- metabolic disorder characterized by ovulatory dysfunction, hirsutism, orelevated levels of androgens in blood. Eleven PBDEs were analyzed in blood, but only six had detectableserum concentrations (however, only five are listed in the report: BDE 28, BDE 47, BDE 99, BDE 100, and BDE 153). Cases and controls had comparable serum concentrations of these BDEs, and in both groups, BDE 47 had the highest concentration (~25.5 ng/g lipid). After controlling for age, BMI, and ethnicity, none of the measured BDEs was associated with increased risk of PCOS. PBDEs 1893. HEALTH EFFECTSConcentrations of PBDEs in omental fat or serum were not associated with risk of uterine fibroids in women undergoing laparoscopy or laparotomy in a study conducted in Salt Lake City, Utah, and San Francisco, California (Trabert et al. 2015). Ninety-nine women had fibroids and 374 had none. Seven PBDEs were measured: BDE 47, BDE 99, BDE 100, BDE 153, BDE 154, BDE 183, and BDE 209. Whether in omental fat or serum, geometric mean concentrations of PBDEs were comparable betweenwomen who had fibroids and those who did not. Odds of a fibroid diagnosis (and corresponding 95%CIs) showed no association with PBDEs.Acute-Duration Animal StudiesLower-brominated PBDEs: There are limited data regarding the reproductive effects of acute exposure tolower-brominated PBDEs in females. In rat dams exposed once to tetraBDE at 0, 0.14, or 0.7 mg/kg on GD 6 via gavage, a significant 17% increase in paired ovary weight was observed at 0.14 mg/kg, but not0.7 mg/kg (Talsness et al. 2008). No treatment-related changes in gravid uterine weight or number ofimplantation or resorptions were observed in rat dams exposed to pentaBDE at doses up to 2 mg/kg/dayfrom GD 6 to 19 (Blanco et al. 2012). In ovariectomized female mice, no exposure-related changes were observed in uterine wet weight, uterine epithelial height, or vaginal epithelium thickness followingexposure to pentaBDE at doses up to 300 mg/kg/day for 3 days via gavage, with or without co-exposureto β-estradiol-3-benzoate (Mercado-Feliciano and Bigsby 2008a). The effects of acute pentaBDE exposure on androgen-dependent tissue weights was assessed in castratedmale rats supplemented with subcutaneous injections of testosterone (to ensure equal levels of circulatingtestosterone in the exposed and control groups) in a Herschberger assay (Stoker et al. 2005). Male ratsthat were castrated on PND 42 were exposed to pentaBDE at 0, 30, 60, 120, or 240 mg/kg/day fromPND 53 to 61 via gavage, and daily gavage exposures were followed by daily subcutaneous injections of0.4 mg/kg testosterone. On PND 62, significant decreases in androgen-dependent tissue weights includeda ~20–55% decrease in ventral prostate weight at ≥30 mg/kg/day, a ~20–60% decrease in seminal vesicle weight at ≥60 mg/kg/day, a ~28–41% decrease in Cowper’s gland weight at ≥120 mg/kg/day, and a 22– 29% decrease in the gland penis and levator ani bulbo cavernosus weight at 240 mg/kg/day (Stoker et al. 2005). There is limited evidence for exposure-related effects on serum reproductive hormones levels in male rats following acute exposure to pentaBDE. Serum testosterone levels were significantly decreased by ~40– 45% in male rats 45 days after a single gavage exposure to pentaBDE at doses ≥0.6 mg/kg; serum PBDEs 1903. HEALTH EFFECTSprogesterone was significantly decreased by ~40% at 1.2 mg/kg (data presented graphically) (Alonso etal. 2010). Significantly increased serum levels of LH were reported in male rats (~65%, data reported graphically) following exposure to pentaBDE at 60 mg/kg/day for 3 days (Stoker et al. 2005). No exposure-related changes were observed in serum testosterone, androsteridione, or estrone levels at doses up to 60 mg/kg/day (Stoker et al. 2005). Following a single gavage administration of pentaBDE at 0, 0.06, or 0.3 mg/kg on GD 6, no exposure-related changes in serum testosterone or LH levels were observed in F1 male rats on PND 140 (Kuriyama et al. 2005). As discussed in Section 3.2.2.6 (Developmental Effects), F1 reproductive success was assessed followinga single gavage administration of pentaBDE at 0, 0.06 or 0.3 mg/kg or tetraBDE at 0, 0.14, or 0.7 mg/kgon GD 6 in rat dams (Kuriyama et al. 2005; Talsness et al. 2005, 2008). Following pentaBDE exposure, no significant exposure-related effects were observed in F1 male fertility when exposed males were matedwith unexposed females, and the only mating behavior significantly altered was a 32% decrease in thepercent of males with two or more ejaculations in F1 males from the 0.3 mg/kg group (Kuriyama et al. 2005). Similarly, no changes in female pregnancy rate was observed in F1 females mated to unexposed males following exposure to pentaBDE or tetraBDE (Talsness et al. 2005, 2008). F1 male reproductiveperformance was not assessed following tetraBDE exposure. Despite a lack of exposure-relatedimpairment in reproductive success following acute developmental exposure to penta- or tetraBDE, reductions in testicular weight, sperm/spermatid number, and daily sperm production were observed in F1males exposed to pentaBDE at ≥0.06 mg/kg/day, and a decreased number of secondary follicles andultrastructural changes in the ovaries were observed in F1 females exposed to pentaBDE at≥0.06 mg/kg/day or tetraBDE at ≥0.14 mg/kg/day (Kuriyama et al. 2005; Talsness et al. 2005, 2008). In a one-generation study in mink, females exposed to pentaBDE at doses ≥0.25 mg/kg/day from premating day 28 through PNW 6 did not whelp (Bull et al. 2007; Zhang et al. 2009). It is not clear in theBull et al. (2007) study whether mink exposed to 0.25 mg/kg/day never became pregnant or had completelitter loss. However, Zhang et al. (2009) reported that females exposed to 0.31 mg/kg/day had no exposure-related changes in mating success; rather, sows showed complete litter loss with 70% showingclear postimplantation loss.DecaBDE: In mouse dams exposed to decaBDE at 0, 150, 750, 1,500, or 2,000 mg/kg/day via gavagefrom GD 7 to 9, the percentage of postimplantation loss per litter was significantly increased by 3, 2.7, and 9.8% at 750, 1,500, and 2,000 mg/kg/day, respectively, compared with control (Chi et al. 2011). At1,500 and 2,000 mg/kg/day, the percentage of resorptions per litter was also significantly increased by PBDEs 1913. HEALTH EFFECTS2.7 and 8.6%, respectively. Additionally, the percentage of live fetuses per litter was significantlydecreased by 10% in the high-dose group (Chi et al. 2011). These effects could reflect reproductivetoxicity in the dams or developmental toxicity in the fetuses (see Section 3.2.2.6, Developmental Effects).No other acute-duration studies analyzing reproductive effects were located for decaBDE.Intermediate-Duration Animal StudiesLower-brominated PBDEs: No exposure-related changes were observed in reproductive end points(number of pregnancies, gestation length, number, size, or sex ratio of litters) in rats exposed to pentaBDE at doses up to 25 mg/kg/day for 70 days prior to mating until PND 42 via gavage (Bondy et al. 2013), in rats fed vanilla wafers containing pentaBDE at doses up to 11.4 mg/kg/day from pre-matingday 28 through PND 21 (Poon et al. 2011), in mice exposed to tetraBDE at doses up to 0.1 mg/kg/day viagavage from pre-mating day 28 through PND 21 (Woods et al. 2012), or in mice fed cornflakescontaining tetraBDE at doses up to 1 mg/kg/day from pre-mating day 28 through PND 21 (Koenig et al. 2012; Ta et al. 2011). Similarly, no exposure-related effects on litter parameters (successful delivery oflitters, gestation length, litter size, sex ratio, number of live pups) were observed in rats or mice exposed to pentaBDE at doses up to 2 or 10 mg/kg/day, respectively, from GD 1 or 6 to PND 21 via gavage(Branchi et al. 2005; Cheng et al. 2009; Zhao et al. 2014), in rats exposed to pentaBDE at doses up to 30 mg/kg/day from GD 1 to PND 21 via dosed cookies (Bowers et al. 2015), or rats exposed to dietarytetraBDE at doses up to 32 mg/kg/day from GD 1 to PND 14 (Wang et al. 2011a). The number of litterssurviving until PND 8 was significantly decreased following exposure to tetraBDE at 0.1 mg/kg/day frompre-mating day 28 through PND 21 in one study (Woods et al. 2012); however, reduced pup survival wasnot reported in other studies (Bondy et al. 2013; Koenig et al. 2012; Poon et al. 2011; Ta et al. 2011;Wang et al. 2011a). Male rats exposed to tetraBDE at 1 mg/kg/day via gavage for 8 weeks showed a significant 24% decrease in daily sperm production; no exposure-related effects were observed at doses ≤0.03 mg/kg/day (Zhang etal. 2013b). In another study, sperm morphology, motility, and capacitation were evaluated in micefollowing gavage exposure to tetraBDE at 0, 0.0015, 0.045, 1.5, or 30 mg/kg/day for 30 days (Wang et al. 2013). No exposure-related changes in sperm morphology or sperm motility were observed. Asignificantly decreased rate of sperm capacitation (% B-type [mature] sperm) was observed in the 0.0015, 0.045, and 30 mg/kg/day groups, but not in the 1.5 mg/kg/day group (Wang et al. 2013). No exposure-related changes in sperm counts, motility, or DNA damage were observed in rats exposed to a penta PBDEs 1923. HEALTH EFFECTSdecaBDE mixture (52.1% pentaBDE, 44.2% decaBDE, 0.4% octaBDE) at dietary doses up to20 mg/kg/day for 70 days (Ernest et al. 2012).Microscopic changes in the testes of rats were observed following exposure to tetraBDE at≥0.03 mg/kg/day via gavage for 8 weeks, including increased epithelial thickness, vacuolar spaces in the seminiferous epithelium, and increased number of multinucleated giant cells (arising from spermatocytes that aborted meiosis); no histopathological changes were observed at 0.001 mg/kg/day (Huang et al. 2015;Zhang et al. 2013b). Additionally, the number of apoptotic cells was significantly increased by 2- and 3-fold in the testes of rats from the 0.03 and 1 mg/kg/day groups, respectively (Huang et al. 2015; Zhanget al. 2013b). In similarly-exposed rats, co-treatment with dexamethasone phosphate (DEX; a CYP3A1 inducer) “aggravated” the observed histopathological effects (Zhang et al. 2013b). Mice exposed to tetraBDE at 0, 0.0015, 0.045, 1.5, or 30 mg/kg/day via gavage for 30 days also showed testicular effectsat doses ≥0.045 mg/kg/day, including “some” seminiferous tubules with complete germ cell loss and a Sertoli cell-only phenotype (incidence not reported) and a dose-related increase in the number ofapoptotic cells (Wang et al. 2013). No exposure-related changes were observed in testes weight at doses up to 30 mg/kg/day (Wang et al. 2013). In other studies, exposure-related changes were not observed inorgan weight or histology in male or female reproductive tissues from mice exposed to tetraBDE atdietary doses of 0.45 mg/kg/day for 28 days (Maranghi et al. 2013), rats exposed to penta- or octaBDE atdietary doses up to 750 mg/kg/day for 28–90 days (IRDC 1976, 1977; WIL Research Laboratories), ratsexposed to a penta-decaBDE mixture (52.1% pentaBDE, 44.2% decaBDE, 0.4% octaBDE) at dietarydoses up to 20 mg/kg/day for 70 days (Ernest et al. 2012), rats exposed to pentaBDE at gavage doses upto 250 mg/kg/day for 15–28 days (Becker et al. 2012; Oberg et al. 2010), or ovariectomized mice exposed to pentaBDE at 50 mg/kg/day for 34 days via gavage (with or without β-estradiol-3-benzoate co-exposure) (Mercado-Feliciano and Bigsby 2008a). Additionally, organ weight changes were not observedin F0 or F1 rats exposed to pentaBDE at doses up to 25 mg/kg/day via gavage from pre-mating day 70 toPND 42 (Bondy et al. 2013) or F0 or F1 mink exposed to dietary pentaBDE at doses up to 0.31 mg/kg/day for 4 weeks prior to mating through PNW 6 or 33 (Zhang et al. 2009).Following exposure to pentaBDE at 0, 3, 30, or 60 mg/kg/day for 15 days via gavage, serum prolactin was significantly decreased by 67% at 60 mg/kg/day in male rats (Becker et al. 2012). While testosterone and FSH levels were not significantly altered at any specific dose (based on pair-wise analysis), trendtests showed significant dose-dependent increases in testosterone and FSH levels; LH and E2 were belowthe detection limit in all groups (Becker et al. 2012). In contrast, a repeat of the same study in a differentlaboratory showed no exposure-related changes in serum testosterone, LH, FSH, E2, or prolactin in male PBDEs 1933. HEALTH EFFECTSrats exposed to pentaBDE at doses up to 60 mg/kg/day for 15 days (Becker et al. 2012). Similarly, malerats exposed to a penta-decaBDE mixture (52.1% pentaBDE, 44.2% decaBDE, 0.4% octaBDE) at dietarydoses up to 20 mg/kg/day for 70 days did not show altered serum testosterone levels (Ernest et al. 2012).After exposure to tetraBDE for 8 weeks via gavage (6 days/week), male rats showed significantreductions in serum testosterone of ~35, 54, and 63% at 0.001, 0.03, or 1 mg/kg/day, respectively (data reported graphically) (Zhang et al. 2013b). No exposure-related changes were observed in serum E2,FSH, or LH levels at doses up to 1 mg/kg/day (Zhang et al. 2013b). In female mice exposed to dietarytetraBDE for 28 days, serum testosterone and E2 were significantly increased by 57 and 18% at0.45 mg/kg/day (only tested dose) (Maranghi et al. 2013).Biochemical analysis of rat testes showed significant dose-related elevations of mRNA levels of severalapoptosis genes following exposure to tetraBDE at doses of 0.001–1 mg/kg/day via gavage for 8 weeks(6 days/week) (Zhang et al. 2013b). Additionally, elevated levels of ROS were observed at 1 mg/kg/day(Zhang et al. 2013b). Co-exposure to DEX (a CYP3A1 inducer) enhanced ROS-induction in the testes, with significant elevations observed at ≥0.001 mg/kg/day (Zhang et al. 2013b). No exposure-relatedchanges were observed in testicular mRNA expression levels for genes involved in steroidogenesis (Star,Cyp17a1, Ar, Srd5a1, Srd 5a2, Cyp19a1, Esr1, Esr2) in rats following exposure to a penta-decaBDEmixture (52.1% pentaBDE, 44.2% decaBDE, 0.4% octaBDE) at dietary doses up to 20 mg/kg/day for70 days (Ernest et al. 2012).Exposure to lower-brominated PBDEs has been reported to cause reproductive effects in developing rats.Two companion studies evaluated reproductive system development following exposure to pentaBDE at0, 3, 30, or 60 mg/kg/day via gavage during pubertal development (PNDs 23–53 in males or PNDs 22–41 in females) (Stoker et al. 2004) or 0, 60 or 120 mg/kg/day via gavage during pubertal development in males (PNDs 23–53) (Stoker et al. 2005). In males, preputial separation (PPS) was significantly delayed by 1.7–5 days at ≥30 mg/kg/day, absolute ventral and lateral prostate weights and seminal vesicle weights were significantly decreased 16–29% at ≥60 mg/kg/day (in the absence of body weight effects), and serum prolactin was increased 2-fold at 60 mg/kg/day (not evaluated in the 120 mg/kg/day group). Therewere no exposure-related changes in testicular weight or histology, epididymides histology, serumtestosterone or LH, or pituitary LH or prolactin levels at doses up to 60–120 mg/kg/day (Stoker et al. 2004, 2005). In females, vaginal opening was significantly delayed by 1.8 days in the 60 mg/kg/daygroup. No changes were observed in estrous cycling or ovarian or uterine weight or histology at doses up to 60 mg/kg/day (Stoker et al. 2004). As summarized in Section 3.2.2.6 Developmental Effects, observed effects in F1 rats exposed to pentaBDE at 30.1 mg/kg/day from GD 6 to PND 21 included decreased PBDEs 1943. HEALTH EFFECTSanogenital distance, delayed PPS, and decreased serum testosterone in male offspring and decreasedmammary gland development in female offspring (Kodavanti et al. 2010).DecaBDE: Information on effects of intermediate-duration exposure to decaBDE on reproductivefunction is limited to negative findings in a one-generation study in rats using a low-purity (77.4%)decaBDE mixture (Dow Chemical Co. 1975; Norris et al. 1975a). Male and female rats were exposed to0, 3, 30, or 100 mg/kg/day doses in the diet for 60 days prior to mating through PND 21. Parametersmonitored included length of time between first day of cohabitation and parturition, numbers of live and dead newborn, number of live pups (PNDs 1, 7, 14, and 21), litter weight (PNDs 1, 7, and 14), and weanling weight (PND 21). Comprehensive histological examinations (adults and weanlings), skeletalexaminations (weanlings), and cytogenetic evaluation of bone marrow (adults and weanlings) were also performed on PND 21. There were no exposure-related effects on reproductive parameters or anyindications of maternal or neonatal toxicity. Additionally, no exposure-related changes were observed inlitter parameters in mice exposed to decaBDE at gavage doses up to 1,500 mg/kg/day from GD 0 to 17 (Tseng et al. 2008), in rats exposed to decaBDE at gavage doses up to 1,000 mg/kg/day from GD 6 toPND 21 (Biesemeier et al. 2011), or in rats or mice exposed to decaBDE at dietary doses up to 146 or3,100 mg/kg/day, respectively, from GD 10 to PND 21 (Fujimoto et al. 2011; Watanabe et al. 2008, 2010b).Sperm parameters have been assessed in rats and mice following intermediate-duration exposure to decaBDE. In the rat study, no exposure-related changes were observed in male epididymal sperm counts or morphology following exposure to decaBDE at gavage doses up to 60 mg/kg/day for 28 days (Van derven et al. 2008a). In one mouse study, sperm count and viability were significantly reduced by about40% in males exposed to 950 mg/kg/day via gavage for 35 days, compared with controls; no exposure-related changes in sperm count and viability were observed at 750 mg/kg/day (Sarkar et al. 2015). In contrast, no exposure-related changes were observed in sperm count, sperm motility, or the percent ofabnormal sperm heads in mice exposed to decaBDE at doses up to 1,500 mg/kg/day via gavage for50 days (Tseng et al. 2006). For sperm velocity, the lateral head amplitude was significantly decreased inthe 500 and 1,500 mg/kg/day groups, compared with control; no exposure-related changes were observedin curvilinear, average path, or straight line velocity or beat-cross frequency. No exposure-relatedchanges were observed in sperm damage parameters; however, sperm H2O2 production was significantlyincreased at 500 and 1,500 mg/kg/day (no change in O2- production). The percentage of sperm with high mitochondrial membrane potential was significantly decreased in the 1,500 mg/kg/day group, compared with controls. PBDEs 1953. HEALTH EFFECTSThere is limited evidence of histopathological damage in reproductive organs following intermediate-duration exposure to decaBDE. Various degenerative changes were observed in the seminiferous tubules of mice exposed to decaBDE at gavage doses of 950 mg/kg/day for 35 days, including thinning of thegerminal epithelium, decreased diameter of the seminiferous tubules, depletion of germ cells, exfoliationof germ cells, and intraepithelial vacuolation (Sarkar et al. 2015). Both the height of the germinal epithelium and the diameter of seminiferous tubules were significantly decreased at 950 mg/kg/day, compared with controls. No exposure-related testicular lesions were observed at 750 mg/kg/day. Infemale rats exposed to decaBDE at 0 or 300 mg/kg/day from 3 weeks of age, through mating to untreated males, gestation, and lactation (~11 weeks), the ovaries in the exposed rats had significantly increased“histological scores” than the ovaries from the control rats (methods of histological scoring and incidences of lesions were not reported) (Liu et al. 2012). Observations in the ovaries of exposed ratsincluded atrophic changes, decreased number of follicles, and increased fibrotic tissue (Liu et al. 2012). F0 reproductive success and F1 developmental end points were not reported in this study. However, in other intermediate-duration studies, no histopathological changes were observed following exposure to decaBDE in male or female reproductive tissues from rats exposed for 28–50 days at gavage doses up to1,500 mg/kg/day (Tseng et al. 2006; Van der ven et al. 2008a), rats exposed for 28–30 days at dietarydoses up to 800 mg/kg/day (IRDC 1976; Norris et al. 1973, 1975a), or rats or mice exposed for 13 weeksat dietary doses up to 8,000 or 9,500 mg/kg/day, respectively (NTP 1986).In a 28-day study, rats were exposed to decaBDE at 0, 1.7, 3.75, 7.5, 15, 30, or 60 mg/kg/day via gavage(Van der ven et al. 2008a). The study authors reported a dose-related decrease in epididymis weight inmale rats (maximal decrease of 22.5%) and a dose-related increase in seminal vesicle/coagulation glandweight (maximal increase of 38.3%); however, the lowest doses at which the effects were observed were not reported. Instead, results were reported in terms of BMD analysis (BMDRD10% for epididymis=4.0 mg/kg/day, BMDLRD10% was not determined; BMD/BMDLRD10% for seminal vesicle=1.5/0.2 mg/kg/day). No exposure-related changes in organ weight were reported for testes, ovaries, oruterus at doses up to 60 mg/kg/day (Van der ven et al. 2008a). In other rat studies, no exposure-relatedchanges in testes, epididymides, or seminal vesicle weights were observed following exposure to dietarydecaBDE doses up to 800 mg/kg/day for 28–30 days (IRDC 1976; Norris et al. 1973, 1975a). In mice, one study reports a significant 13–18% decrease in relative testis and epididymides weight followingexposure to decaBDE at 950 mg/kg/day via gavage for 35 days, compared with controls; no exposure-related changes in male reproductive organ weights were observed at 750 mg/kg/day (Sarkar et al. 2015). In contrast, no exposure-related changes in testes, epididymides, or seminal vesicle weights were PBDEs 1963. HEALTH EFFECTSobserved in another mouse study that exposed animals to decaBDE at gavage doses up to 1,500 mg/kg/day for 50 days (Tseng et al. 2006). In a 35-day study, serum testosterone levels were significantly decreased by 65% in mice followingexposure to decaBDE at gavage doses of 950 mg/kg/day, compared with controls; testosterone levels at750 mg/kg/day were comparable to control (Sarkar et al. 2015).There is limited evidence that exposure to decaBDE in developing animals results in reproductive effects.In male rats exposed during pubertal development, exposure to decaBDE at doses up to 600 mg/kg/day did not cause exposure-related changes in testes, epididmides, or prostate organ weight, testicularhistology, or testicular mRNA expression levels of steriodogensis-related genes (Lee et al. 2010). As summarized in Section 3.2.2.6 (Developmental Effects), only one intermediate-duration gestationalexposure study reported reproductive effects in mouse offspring exposed to decaBDE doses of 10– 1,500 mg/kg/day from GD 0 to 17 via gavage, including testicular lesions, decreased AGD, and alteredsperm parameters (Tseng et al. 2013). No exposure-related changes in AGD, onset of puberty (PPS orvaginal opening), estrous parameters, or reproductive organ weight and histology were reported in offspring exposed to decaBDE at doses up to 1,000 mg/kg/day during gestation and lactation (Biesemeieret al. 2011; Fujimoto et al. 2011) or doses up to 20 mg/kg/day from PND 2 to 15 (Rice et al. 2007).Chronic-Duration Animal StudiesLower-brominated PBDEs: No chronic-duration studies evaluating reproductive system effects were located for lower-brominated PBDEs.DecaBDE: No histopathological changes were observed in male or female reproductive tissues from rats or mice that were exposed to decaBDE at dietary doses up to 2,550 or 7,780 mg/kg/day, respectively, for103 weeks (NTP 1986). In the only other chronic-duration study, a low-purity mixture (77.4% purity) did not cause histopathological changes in male or female reproductive tissues from rats at doses up to1.0 mg/kg/day for 2 years (Kociba et al. 1975; Norris et al. 1975a),Summary. Based on the evidence in humans and animals, lower-brominated PBDEs are potentially toxic to the male reproductive system in humans, including the developing reproductive system in children (seeSection 3.2.2.6, Developmental Effects for more details). Available data for decaBDE provide verylimited evidence of male reproductive damage, and are insufficient to determine if oral decaBDE PBDEs 1973. HEALTH EFFECTSexposure can damage the male reproductive system in humans. For female reproductive end points, dataare inconsistent in humans and animals; therefore, it is unclear whether PBDEs affect the female reproductive system in adults or developing infants/children. The highest NOAEL values and all LOAELvalues from each reliable study for reproductive effects in each species and duration category are recorded in Table 3-2 and plotted in Figure 3-2.3.2.2.6 Developmental Effects Human Studies. Numerous studies have investigated potential relationships between developmentalPBDE exposure and birth outcomes (e.g., birth weight, length, head circumference, gestational age) and/or postnatal development of various systems (neurological, reproductive, endocrine, immune).Physical Growth and Development and Related Birth OutcomesRobledo et al. (2015a) studied potential relationships between pre-conception maternal and paternalserum levels of 10 PBDEs (BDE 17, BDE 28, BDE 47, BDE 66, BDE 85, BDE 99, BDE 100, BDE 153, BDE 154, and BDE 183) and birth weight and length, head circumference, and ponderal index (100x[birth weight in g/birth length in cm3]) in a prospective cohort study in Michigan and Texas (LIFEstudy). Geometric means of maternal and paternal serum PBDE levels (n=234 pairs) were 0.001– 0.116 ng/g serum and 0.001–0.113 ng/g serum, respectively. The relative concentrations of PBDEs was similar in maternal and paternal serum, with BDE 47 having the highest level (Robledo et al. 2015b). Levels were not reported in terms of ng/g lipid. Maternal serum BDE 28 and BDE 183 levels and paternal serum BDE 183 were significantly associated with lower birth weight in female infants, with adecrease of 84.6–151.33 g per 1 standard deviation (SD) increase in ln- transformed chemicalconcentration. Maternal serum BDE 28 was also significantly associated with smaller birth length(-1.14 cm per 1-SD increase) and head circumference (-1.05 cm per 1-SD increase) in female infants.However, in male infants, maternal serum BDE 66 and BDE 99 were significantly associated with higherbirth weight (+85.21–125.04 g per 1-SD increase) and larger head circumference (+0.6–0.91 cm per 1-SD increase. Maternal serum BDE 99 was also significantly associated with larger birth length in boy infants (+0.76 cm per 1-SD increase). No other significant associations were observed.Maternal serum concentrations were studied in a population of 286 low-income women living in theSalinas Valley of California (a subset of the CHAMACOS study) (Harley et al. 2011). The main PBDEsdetected in maternal blood, collected during the second trimester of pregnancy, were BDE 47, BDE 99, BDE 100, and BDE 153 (median concentrations of 14.57, 3.85, 2.45 and 2.03 ng/g lipid, respectively). PBDEs 1983. HEALTH EFFECTSSerum concentrations of each of these congeners, and their sum, were significantly inversely related toinfant birth weight in crude analysis. After adjustment for covariates, regression analysis showed thateach 10-fold increase in BDE 47, BDE 99, or BDE 100 was associated with roughly a 115-g decrease inbirth weight. These associations were no longer statistically significant when maternal weight gain was included in the models. There were no significant relationships between maternal serum PBDE and infantbirth length or head circumference, or length of gestation. In a follow-up study of the CHAMACOScohort, the potential association between obesity indices (BMI, waist circumference, obesity/overweightstatus) at 2–7 years of age and maternal serum PBDE levels was evaluated (BDE 47, BDE 99, BDE 100, BDE 153, and their sum); child serum PBDE levels were also determined at age 7. The geometric means of the total maternal (n=224) and child (n=216) serum PBDE concentration were 25.35 and 83.03 ng/glipid, respectively (Erkin-Cakmak et al. 2015). No significant associations were observed betweenobesity measures and maternal serum concentrations of individual or summed PBDEs at any age when both sexes were examined together; however, statistical analysis revealed a significant effect modificationby sex. When evaluated separately, a significant negative relationship was observed between BMIz-score in 3.5-year-old females and a 10-fold increase in summed maternal serum PBDE levels (adjusted β -0.64, 95% CI -1.23, -0.06), and a significant positive relationship was observed between BMI z-score in 3.5-year-old males and a 10-fold increase in summed maternal serum PBDE levels (adjusted β 0.99,95% CI 0.32, 1.66). This sex difference was also observed at 7 years of age, with a significant negativeassociation in females (adjusted β -0.41, 95% CI -0.87, -0.05) and a near-significant positive associationin males (adjusted β 0.26, 95% CI: -0.19, 0.72). Similar trends were observed for waist circumferenceand obesity status (data not reported). When evaluating child serum PBDE levels, significant associations after adjustment for potential cofounders included a negative association between BDE 153 and ∑PBDEsand BMI and waist circumference and a reduced risk for being overweight. There was no significanteffect by sex modification in the analysis of child serum levels. This follow-up study suggests possibleobesogenic effects of in utero PBDE exposure in boys. In contrast, Vuong et al. (2016b) found no significant association between PBDEs (BDE 28, BDE 47, BDE 99, BDE 100, BDE 153, and their sum) measured in maternal serum at 16 weeks of gestation (geometric mean 39.1 ng/g lipid) and height orweight of boys and girls at 1–8 years of age. However, BDE 153 was associated with lower BMI at 2– 8 years, smaller waist circumference at 4–8 years, and lower percent body fat at 8 years. The Vuong et al. (2016b) study comprised 318 mother-child pairs in the Health Outcomes and Measures of theEnvironment (HOME) Study in Cincinnati, Ohio. Agay-Shay et al. (2015) also evaluated the potential association between maternal PBDE exposure and childhood obesity. This study did not find significant associations between maternal PBDE colostrum levels (BDE 47, BDE 99, BDE 100, BDE 153, BDE 154, BDE 209, and their sum) and BMI z-scores or risk of being overweight (BMI >85th percentile) in PBDEs 1993. HEALTH EFFECTS470 7-year-old boys and girls from a Spanish birth cohort study; however, sex-specific analyses were notconducted. Geometric mean concentrations of BDE 47, BDE 99, BDE 100, BDE 153, BDE 154, and BDE 209 in colostrum were 0.5, 0.3, 0.2, 0.7, 0.7,and 0.8 ng/g lipid, respectively.Results from other studies evaluating potential associations between maternal serum PBDE levels duringpregnancy and/or at delivery and birth outcomes are inconsistent between studies and congeners. TheHealthy Pregnancy, Healthy Baby (HPHB) birth cohort in North Carolina also evaluated potentialassociations between third trimester maternal serum PBDE concentrations and birth weight, length, headcircumference, and birth weight percentile for gestational age (Miranda et al. 2015). Serum wasevaluated for 27 PBDE congeners and 6 OH-PBDEs; however, only congeners and metabolites detectedin >50% of the subjects (n=137) were included in the analysis, including BDE 47, BDE 99, BDE 100, BDE 153, 4’-OH-BDE-49, and 6’-OH-BDE-47 (median serum concentrations of 18.87, 5.5, 4.61, 5.65, 0.12, and 0.19 ng/g lipid, respectively). The only significant finding was a negative association between maternal BDE 153 and head circumference (0.32 cm decrease per 2-fold increase in BDE 153); however, this association was no longer significant after adjustment for maternal risk factors. Similarly, Serme-Gbedo et al. (2016) did not find a statistically significant relationship between birth weight and BDE 47, BDE 99, BDE 100, BDE 153, or their sum in maternal serum collected during early pregnancy(~12 weeks; median concentration of 32.99 ng/g lipid for total PBDE). This study evaluated 349 Canadian women (GESTE birth cohort), and utilized both unadjusted models and multivariateregression models adjusted for a full range of clinical risk factors known to affect fetal growth as well as other environmental pollutants that are likely to impact fetal growth (PCBs, mercury, lead, cadmium, manganese). In contrast, Chen et al. (2015) reported negative associations between lower levels ofmaternal serum PBDEs and birth weight and length in the prospective Laizhou Wan birth cohort (LWBC)in 215 Chinese women. In this study, median maternal serum levels of PBDEs, collected at delivery,were 2.27, 2.26, 3.53, 2.13, and 4.87 ng/g lipid for BDE 28, BDE 47, BDE 99, BDE 100, and BDE 153, respectively (median concentration of total PBDEs was 21.68 ng/g lipid). After adjustment for potential covariants, a near-significant negative association was observed between birth weight and maternal serumBDE 28; when stratified by sex, this decrease was significant for males (253.76 g decreased per 10-fold increase in BDE 28), but not females. A negative association was observed between birth length and maternal serum levels of all congeners; however, the association was only statistically significant forBDE 28 and BDE 100 when males and females were combined (0.92–0.97 cm decrease per 10-fold BDEincrease). When stratified by sex, BDE 99 was significantly negatively associated with male infant lengthand BDE 100 was significantly negatively associated with female infant length (1.47–1.50 cm decrease per 10-fold BDE increase). There were no significant relationships between maternal serum PBDE and PBDEs 2003. HEALTH EFFECTSinfant head circumference or length of gestation. In a case-control study of 197 full-term births and 82 pre-term births, identified at the Centennial Women’s Hospital in Nashville, serum BDE 47 levelswere divided into the following ordinal scale: 25–135, 136–200, 200–321, 322–1,000, and >1,000 pg/mL(Peltier et al. 2015). The distribution for controls into the five bins was 23, 22, 21, 22, and 12%, respectively, while the distribution for cases was 10, 6, 20, 35, and 29%, respectively; these distributionsare statistically significantly different. When cases and controls were combined for analysis, women fromthe two higher exposure bins (>323 pg/mL BDE 47) showed a significantly increased risk for pretermbirth.Several investigators have also evaluated associations between serum maternal and umbilical cord PBDElevels, and their potential associations with birth indices. As observed in maternal serum studies, data are inconsistent between studies and congeners. In a study of 97 Canadian women (a subset of the FAMILY study), Foster et al. (2011) found no significant relationship between infant birth weight and maternalserum PBDE (BDE 17, BDE 28, BDE 47, BDE 66, BDE 99, BDE 100, BDE 153, BDE 154, BDE 183, and their sum), whether the serum was collected during the second trimester of pregnancy or at delivery.PBDE concentrations in this study were relatively high (median concentrations of 52.1 and 50.1 ng/g lipid for total PBDE at mid-pregnancy and delivery). Foster et al. (2011) also evaluated the potentialrelationship between umbilical cord serum PBDE (BDE 17, BDE 28, BDE 47, BDE 66, BDE 99, BDE 100, BDE 153, BDE 154, BDE 183, and their sum) and birth weight. They found a significantnegative association between birth weight and concentrations of one PBDE congener (BDE 99) in umbilical cord serum. PBDE concentrations in umbilical cord serum were 1.7–3.4 times higher than in maternal serum at delivery (median concentration of 100.0 ng/g lipid for total PBDE). These results are in contrast to Mazdai et al. (2003), who found umbilical cord concentrations of PBDE (BDE 47, BDE 99, BDE 100, BDE 153, BDE 154, BDE 183, and their sum) to be similar to maternal serum concentrationsat delivery (median concentrations of 39 and 37 ng/g lipid for total PBDE, respectively) in a small studyof 12 women recruited upon presenting in labor to Indianapolis hospitals. In this study, no relationship ofmaternal or cord blood serum to infant birth weight was observed. In contrast to these findings, a lowexposure study in 686 Spanish women (INMA cohort; median total serum PBDE concentrations in maternal blood collected at gestational week 12 and umbilical cords of 10.74 and 7.51 ng/g lipid, respectively) reported a significant relationships between PBDE levels and some birth indices (Lopez-Espinosa et al. 2015). While concentrations were similar between maternal and umbilical cords serum,Pearson’s correlations between maternal and cord log2(PBDEs) were low (0.18, 0.06, 0.04, 0.09, and 0.07 for BDE 47, BDE 99, BDE 153, BDE 159 and BDE 209, respectively). Potential associations wereexamined between maternal or cord serum and birth weight, length, and head circumference for each PBDEs 2013. HEALTH EFFECTScongener and their sum. A significant inverse association was observed between birth weight and maternal serum BDE 99 levels (Lopez-Espinosa et al. 2015). After adjustment for covariates, regression analysis showed that each 2-fold increase in BDE 99 was associated a 1.4% decrease (~46.8 g) decreasein birth weight. Maternal serum BDE 99, as well as the sum of all PBDEs, was also significantlyassociated with a 2.1–2.9% decrease in head circumference per 2-fold increase in PBDE concentration (~0.7–1.0 cm). No significant associations were observed between birth weight or head circumference for other maternal serum congeners or any cord serum congeners, and no significant associations were observed for birth length. Lopez-Espinosa et al. (2015) also evaluated fetal growth indices duringpregnancy using ultrasound measurements in mid-pregnancy (gestational weeks 12–20) and latepregnancy (gestational weeks 20–34), including abdominal circumference, fetal length, biparietaldistance, and estimated fetal weight. Significant associations, observed only during gestational weeks20–34, included inverse relationships between maternal BDE 99 and biparietal distance, abdominalcircumference, and estimated fetal weight; cord BDE 99 and ∑PBDEs and fetal abdominal circumferenceand estimated fetal weight; and maternal ∑PBDEs and biparietal distance (1.3–3.5% decrease per 2-fold increase in PBDE concentration). In another low-exposure study, Tan et al. (2009) also found no relationship between PBDE in umbilical cord blood and infant birth weight (or length, head circumference, or sex) in babies from 41 native mothers admitted to the Singapore National Hospital forCesarean section. Median PBDE concentrations in this study population 3.3 ng/g lipid for total PBDE in cord blood. There was reported to be a small positive association between cord blood concentrations of the PBDE congeners BDE 47 and BDE 99 and Apgar score at 1 minute in this study. Wu et al. (2010) also studied cord blood, but did not assess effect on birth weight alone, using instead acomposite of adverse birth outcomes, including low birth weight but also premature delivery and stillbirth. These researchers reported significant associations between umbilical cord PBDE (BDE 28, BDE 47, BDE 99, BDE 153, BDE 183, and total) and adverse birth outcomes in a comparison of128 normal births and 25 cases of adverse birth outcomes from two towns (Guiyu and Chaonan) in China. The two towns represent an e-waste recycling area (Guiyu, with a median total PBDE concentration of13.8 ng/g lipid [n=102]) and a control area with no e-waste recycling workshops (Chaonan, with a mediantotal PBDE concentration of 5.2 ng/g lipid [n=51]). In a companion study, neonatal physiological indiceswere compared between 69 births in Guiyu (e-waste recycling) and 86 births in Haojiang (no e-waste re cycling), using placental PBDE concentrations as an exposure metric (Xu et al. 2015b). Median totalplacental PBDE concentrations were 32.25 ng/g lipid in Guiyu and 5.13 ng/g lipid in Haojiang; measuredcongeners included BDE 29, BDE 47, BDE 99, BDE 100, BDE 153, BDE 154, BDE 183, and BDE 209 (mean placental concentrations of 11.66, 1.61, 0.32, 0.14, 3.17, 0.34, 2.20, and 3.30 ng/g lipid, PBDEs 2023. HEALTH EFFECTSrespectively). Head circumference, BMI, and Apgar1 score were significantly decreased by 4–8% and body length significantly increased by 4% in the Guiyu cohort, compared with the Haojiang cohort. No differences were observed in birth weight between the two cohorts. When data from both cohorts werecombined for multiple linear regression analysis, significant negative associations were observed between∑PBDEs and BDE 47 and head circumference; ∑PBDEs, BDE 47, and BDE 99 and neonatal BMI, and ∑PBDEs, BDE 29, BDE 47, BDE 153, and BDE 183 and Apgar1 score. A significant positiveassociation was observed between BDE 47 and body length. Concentrations of PBDE in breast milk have also been studied in relation to adverse birth outcomes. Astudy of births from 20 healthy pregnant women in Taiwan found that increased PBDE (BDE 47, BDE99, BDE 100, and BDE 209) in breast milk was associated with significantly reduced infant birth weight, length, chest circumference, and Quetelet’s index (BMI) (Chao et al. 2007). The median total PBDEconcentration in breast milk was 3.65 ng/g lipid in the study population. A significant negativeassociation between breast milk PBDE concentrations (BDE 47, BDE 99, BDE 100, BDE 153, and theirsum) and birth weight was also identified in a Swedish cohort of 254 women with a median total PBDEconcentration of 2.4 ng/g lipid (Lignell et al. 2013). In contrast, a significant positive association was observed between colostrum (breast milk collected in the first week after delivery) PBDE concentrations (BDE 47, BDE 99, BDE 100, BDE 153, and their sum) and birth weight and birth length in a Tanzanian cohort of 95 women with a median total PBDE concentration of 19.8 ng/g lipid (Müller et al. 2016). When stratified by sex, the only significant finding was a positive association between birth weight and PBDE concentration in female infants. The high PBDE levels were attributed to the consumption ofPemba (clay soil), which is sold in Tanzanian markers to pregnant women as a mineral supplement and nausea-relieving aid. Women who ate Pemba (64%) had significantly higher PBDE concentrations in colostrum, compared to non-Pemba eating women. Reproductive System DevelopmentMain et al. (2007) found a significant positive relationship between concentrations of PBDE in breastmilk and congenital cryptorchidism (undescended testes) in male offspring. The study compared concentrations of 14 PBDE congeners in breast milk of mothers of 62 Danish and Finnish boys with cryptorchidism to mothers of 68 controls from the same population. Significant increases were seen forthe sum of the 7 most prevalent congeners (found in all mothers), for several of the individual congeners, and for the sum of all 14 congeners in cases versus controls. No such relationship was found, however, when placental blood PBDE concentrations, rather than maternal breast milk PBDE concentrations, were PBDEs 2033. HEALTH EFFECTSused to represent exposure. In a comparison of 86 paired samples from this study, PBDE concentrationsin placenta and breast milk were correlated, but absolute PBDE concentrations were 3–4 times higher in breast milk (median of 3.23 ng/g fat for total PBDE) than in placenta (median of 1.19 ng/g fat for totalPBDE). The researchers hypothesized that while breast milk PBDE concentrations reflect theaccumulated body burden of the mother, placental PBDE concentrations may resemble measurements insingle blood samples, reflecting the situation at delivery, but not the long-term exposure. In another case-control study in Danish and Finnish boys, no significant difference was observed in the concentrations of14 PBDE congeners in subcutaneous tissue samples from the inguinal region during orichidopexy surgery(44 cases) or hernia surgery (38 controls) (Koskenniemi et al. 2015). Mean age at surgery was 2.3 and 2.9 years for cases and controls, respectively. Median concentrations of total PBDEs in adipose tissue were 4.90 ng/g in cases and 5.54 ng/g in controls.Other studies of reproductive development found no relationship between PBDE concentrations in mid-pregnancy serum samples from California mothers (median ∑PBDE=33.7 ng/g lipid for 20 cases and 38.6 ng/g lipid for 28 controls) and hypospadias in their male offspring (Carmichael et al. 2010) and no relationship between current serum concentrations of PBDE (range of 4.9–73.6 ng/g lipid, with median of9.9 ng/g lipid) and various measures of sexual maturation (e.g., initiation of breast development, currentbreast development, age at menarche), in a small cohort of 18 (9 boys and 9 girls) teen-aged Dutch children (Leijs et al. 2008). A study of 55 Dutch boys found significant positive associations between maternal serum concentrations of BDE 154 (median=0.5 ng/g lipid) collected on week 35 of pregnancy(but not other congeners measured, including BDE 47 and BDE 153 that occurred at higherconcentrations [medians of 0.9 and 1.6 ng/g lipid, respectively]) and serum concentrations of the sex hormones, E2, free E2, and inhibin B (but not testosterone, LH, FSH, or sex hormone binding globulin),in the baby boys at 3 months of age and testes volume in the boys at 18 months of age, but no effect on penile length at either age (Meijer et al. 2012). In a French birth cohort (n=262 mother-child pairs; 141 male infants, 141 female infants), cord blood total testosterone was significantly decreased in male infants with detectable BDE 209 levels (>0.05 ng/g lipid) compared with male infants without detectableBDE 209 levels (Warembourg et al. 2016). No associations were observed for female total testosteroneor other sex hormones measured in male and/or female infants (sex hormone-binding globulin, E2, freetestosterone, aromatase index, or Anti-Müllerian hormone).Delayed onset of puberty was significantly associated with higher PBDE concentrations in serum in a cohort of 645 ethnically diverse girls (recruited at 6–8 years of age and followed annually) fromCalifornia and Ohio (Breast Cancer and the Environment Research program [BCERP] cohort) (Windham PBDEs 2043. HEALTH EFFECTSet al. 2015a). A cross-sectional analysis, conducted at enrollment, showed a significant decrease in the percentage of females showing Tanner stage 2 breast development per quartile increase in serum ∑PBDE levels; no association was observed with pubic hair development. Median serum ∑PBDE levels atenrollment were 78.3 ng/g lipid (Windham et al. 2015b). In follow-up evaluations, the longitudinal analysis showed a significant increase in age at pubertal transition (both breast and pubic hair Tannerstages) with per quartile increase in serum ∑PBDE levels. When analyzed by individual congener,significant delays in pubic hair development were observed for all congers evaluated (BDE 28, BDE 47, BDE 99, BDE 100, BDE 153, and BDE 154) and significant delays in breast development were observed for BDE 47, BDE 99, BDE 100, and BDE 153 (associations were marginal for BDE 28 and BDE 154).NeurodevelopmentNumerous studies have reported results suggestive of an effect of PBDE on neurodevelopment inchildren. Associations are more consistent when blood concentrations were used as a biomarker ofexposure, compared with studies utilizing breast milk concentrations as a biomarker of exposure.A series of studies evaluated neurodevelopmental outcomes in a longitudinal study of the HOME studycohort from Cincinnati, Ohio (full sample 349 mother-child pairs) (Braun et al. 2014; Chen et al. 2014;Donauer et al. 2015; Vuong et al. 2016a). Several environmental chemicals, including PBDEs, were quantified in maternal blood collected at approximately 16 weeks of gestation. In 5-week-old infants(n=326), no associations were observed between neonatal neurobehavior and maternal serum PBDEs (Donauer et al. 2015). The geometric mean of the sum of the most frequently detected PBDEs (sum4BDE=BDE 47, BDE 99, BDE 100, and BDE 153) in maternal serum from mother-child pairs in theneonatal study was 37.12 ng/g lipid (Donauer et al. 2015). The majority of children from the HOMEcohort (n=309) were examined for cognitive and motor abilities at 1, 2, and 3 years of age, intelligence at5 years of age, and general behavior at 2, 3, 4, and 5 years of age (Behavioral Assessment System forChildren 2 [BASC-2]) (Chen et al. 2014). This analysis focused primarily on maternal BDE 47 exposure(geometric mean=20.1 ng/g lipid; median=18.9 ng/g lipid); however, some analysis were conducted forsum4BDE (geometric mean=37.7 ng/g lipid; median=34.6 ng/g lipid). Maternal serum levels of BDE 47or sum4BDE were not significantly associated with Bayley mental or psychomotor development indices at1–3 years of age. At 5 years, however, adjusted regression analysis showed a significant 4.5-pointdecrease in full-scale IQ per 10-fold increase in maternal serum BDE 47 (95% CI -8.8, -0.1). For thesum4BDE analysis, a marginal decrease in full-scale IQ was observed (β [95% CI] -4.38 [-8.9, 0.14]). Inthe BASC-2 behavioral assessments at 2–5 years of age, significant associations observed included a PBDEs 2053. HEALTH EFFECTSpositive association between maternal BDE 47 serum levels and the Externalizing Problems score at2 and 3 years of age (~2–3-point increase per 10-fold increase in serum BDE 47) and a positive association between maternal BDE 47 serum levels and the Hyperactivity subscore at 2, 3, and 5 years ofage (~2–3-point increase per 10-fold increase in serum BDE 47); other ages showed marginal positiveassociations in these domains. Similar trends were observed in BASC-2 analysis of sum4BDE analysis.Further analysis of neurodevelopment at 4 and 5 years of age in a subset of the HOME cohort (n=175)was reported by Braun et al. (2014), with a specific focus on autistic behaviors (social, repetitive, andstereotypic behaviors). Specific analysis were conducted for BDE 28, BDE 47, BDE 85, BDE 99, BDE100, BDE 153, and BDE 154 (maternal serum geometric mean of 0.8, 20.1, 3.1, 4.7, 3.8, 5.1, and 2.6 ng/glipid, respectively). A marginal positive association was observed between maternal serum BDE 28levels and autistic behaviors (β [95% CI] 2.5 [-0.6, 5.6]). In contrast, a significant negative association was observed between maternal serum BDE 85 levels and autistic behaviors (β [95% CI] -3.2 [-5.9, -0.5]); however, BDE 85 was detected in <50% of the maternal serum samples. In the most recent report,Vuong et al. (2016a) evaluated potential associations between maternal PBDE levels and executivefunction at 5 and/or 8 years in 256 children from the HOME cohort. Specific analysis were conducted forBDE 28, BDE 47, BDE 99, BDE 100, BDE 153, and their sum (maternal serum geometric mean of 1.2, 21.5, 5.0, 4.2, 5.5, and 41.3 ng/g lipid, respectively). After adjusted regression analysis, a significant positive association was observed between impaired behavior regulation and maternal serum BDE 153. Increased BDE 153 was also significantly associated with higher odds of having a behavior regulation (OR, 3.92 95% CI 1.76, 8.73) or a global executive function score (OR 2.34, 95% CI 1.05, 5.23) greaterthan 1 standard deviation above the mean, indicating increased risk of impaired behavior regulation and/or executive function deficits. No significant associations were observed for other congeners or theirsums.In another longitudinal birth cohort of predominantly Mexican-American families in the Salinas Valley ofCalifornia (CHAMACOS), PBDE levels were analyzed in maternal serum (collected at mid-pregnancy orat delivery) and child serum at 7 years of age (Eskenazi et al. 2013). Significant negative associations were found for both maternal serum PBDE and current child serum PBDE and measures of attention, finemotor coordination, and cognitive function (particularly verbal comprehension) in children evaluated at5 (n=249) and/or 7 (n=270) years of age. PBDE concentrations for this subset of the CHAMACOScohort are not available. A follow-up study in the CHAMACOS birth cohort (n=622) evaluated potential associations between PBDE levels in maternal serum (collected at mid-pregnancy or delivery) and currentchild serum (collected at age 9) and numerous measures of attention and executive function at 9, 10.5, and12 years of age (Sagiv et al. 2015). Significant negative associations were observed between several PBDEs 2063. HEALTH EFFECTSmeasures of attention and executive function at 9–12 years of age and maternal serum PBDE (geometric mean of BDE 47, BDE 99, BDE 100, and BDE 153 of 26.3 ng/g lipid), but not current child serum PBDE (geometric mean of BDE 47, BDE 99, BDE 100, and BDE 153 of 63.2 ng/g lipid).Several studies in infants and toddlers report associations between neurodevelopmental deficits andumbilical cord serum PBDE levels. Taiwanese infants (n=36) evaluated for neurological development atage 8–12 months by Bayley Scales of Infant and Toddler Development and parental questionnaire showedsignificant deficits in cognitive score and adaptive behavior associated with umbilical cord bloodconcentrations of total PBDE (range of 2.24–49.1 ng/g lipid, with median of 4.63 ng/g lipid) and concentrations of several individual congeners (Shy et al. 2011). Children from a New York City cohort(n=152) who had higher cord blood concentrations of PBDE congeners BDE 47, BDE 99, and BDE 100 scored significantly lower on tests of mental and physical development at ages 12, 24, and 36 months(Bayley Scales of Infant Development) and ages 48 and 72 months (Wechsler Preschool and PrimaryScale of Intelligence) than children with lower cord blood concentrations (Herbstman et al. 2010). Median and maximum concentrations of BDE congeners in cord blood were 11.2 and 613.1 ng/g lipid forBDE 47, 3.2 and 202.8 ng/g lipid for BDE 99, and 1.4 and 71.9 ng/g lipid for BDE 100. Ding et al.(2015) evaluated the potential association between motor, adaptive, language, and social developmentalquotients in Chinese children at 12 months (n=192) and 24 months (n=149) and cord blood concentrationsof BDE 28, BDE 47, BDE 85, BDE 99, BDE 100, and BDE 153 (median concentrations of 2.05, 3.71, 1.47, 6.70, 2.63, and 2.19 ng/g lipid, respectively) and the sum of BDE 47, BDE 99, BDE 100, and BDE 153 (the four congeners detected in >80% of samples). No significant associations were observed at12 months. At 24 months, a 10-fold increase in cord blood BDE 99 was significantly associated with a2.16-fold decrease in the language domain developmental quotient and a 10-fold increase in cord bloodBDE 47 was significantly associated with a 1.89-fold decrease in the social domain developmentalquotient. No significant associations were observed for the other congeners or the sum of BDE 47, BDE 99, BDE 100, and BDE 153.Studies in preschool and school-aged children (≥4 years of age) also report associations betweenneurodevelopmental impairments and maternal, child, or umbilical cord serum/plasma PBDEs. Both cord and current blood concentrations of BDE 47 (cord blood median and maximum of 2.10 and 16.8 ng/glipid [n=88], current blood median and maximum of 0.12 and 130.2 ng/g lipid [n=244]) were also negatively associated with cognitive and motor functions in 4-year-old children from a Spanish cohort(McCarthy Scales of Children’s Abilities), although in this study, the results were not statisticallysignificant, except for associations with symptoms of poor social competence and ADHD (Gascon et al. PBDEs 2073. HEALTH EFFECTS2011). Positive associations between attention problems (Child Behavior Checklist) and cord plasma levels of BDE 47, BDE 99, BDE 100, and BDE 153 were also observed in 4-year-old children (n=109)from a New York City cohort of mothers pregnant during the September 11, 2001 World Trade Centerterrorist attacks; however, associations were only significant for BDE 47 and BDE 153 (Cowell et al.2015). No significant associations were observed in 6-year-old children. Median BDE 47, BDE 99, BDE 100, and BDE 153 cord plasma levels were 11.2, 3.2, 1.4, and 07 ng/g lipid, respectively.Significant negative correlations were observed for maternal serum PBDE (median ∑PBDE=3.4 ng/g lipid, n=62) in the 35th week of pregnancy and fine manipulative abilities, verbal memory, and sustained attention in a cohort of Dutch children tested at age 5–6 years (Roze et al. 2009), although some aspectsof neurodevelopment appeared to be improved with higher PBDE concentrations in this study, includingcoordination, visual perception, and behavioral outcome reported by parents and teachers. A case-controlstudy of 100 California children, including 51 with autism/autism spectrum, 26 developmentally delayedbut not autistic, and 23 with typical development, found no association of autism or developmental delaywith concentrations of PBDE in serum collected from the children after assessment of developmentalstatus at 36 months (Hertz-Picciotto et al. 2011). PBDE concentrations for this study were not available numerically. Studies attempting to discern relationships between PBDE concentrations in breast milk, rather thanblood, and neurodevelopmental end points produced more uncertain results. A study of 70 infants in Taiwan found no correlation between total PBDE concentrations in breast milk collected within 1 month of delivery (range=1.44–118 ng/g lipid, median=2.92 ng/g lipid) and neurodevelopment as assessed in infants at 8–12 months of age using the Bayley Scales (Chao et al. 2011). There was, however, asignificant inverse association between BDE 209 concentration in breast milk and cognitive score, suggesting delayed cognitive development associated with that particular congener. Gascon et al. (2012)reported similar findings in a Spanish cohort of 290 infants. In this study, total PBDE concentrations in colostrum collected at the hospital within 4 days of delivery (range=0.31–32.66 ng/g lipid, median= 4.05 ng/g lipid) were negatively, but not significantly, associated with the Bayley score for cognitivedevelopment in children assessed at age 12–18 months, but again, a significant negative relationship wasfound for BDE 209, the congener found at the highest concentrations in this population (median of1.02 ng/g lipid). Studies of a North Carolina cohort found modest and imprecisely estimated associationsbetween PBDEs (BDE 28, BDE 47, BDE 99, BDE 100, BDE 153, and their sum) in breast milk collected 3 months postpartum (median ∑PBDE=47.3 ng/g lipid, n=222) and (1) increased extermalizingbehaviors, primarily driven by activity/impulsivity behaviors, in children assessed at 24–36 months forsocial and emotional development; (2) higher anxiety and withdrawal in children assessed at 36 months PBDEs 2083. HEALTH EFFECTSfor behavioral development using a parental rating scale; and (3) improved cognitive skills in childrenassessed at 36 months for cognitive development using the Mullen Scales of Early Learning (Adgent et al. 2014; Hoffman et al. 2012).Potential associations between household BDE 99 and BDE 209 dust and cognitive ability of 6-year-old children (verbal comprehension, working memory) were evaluated in French birth cohort (n=246);potential associations with cord blood BDE 209 were also evaluated (Chevrier et al. 2016). Afteradjustment for potential covariates, children from homes with BDE 99 dust levels above the median value (54 ng/g) or BDE 209 dust levels in the 2nd or 3rd exposure tertile (≥257 ng/g) had significantly increasedrisk of poor performance on verbal comprehension assessments. No association was observed betweendust BDE 99 or BDE 209 levels and working memory, and no association was observed between cord BDE 209 and cognitive function.A case-control study of a Chinese population found no relationship between risk of neural tube defectsand placental PBDE concentrations (median ∑PBDE=0.55 ng/g lipid for 80 cases and 0.54 ng/g lipid for50 controls) (Ma et al. 2012b; Ren et al. 2011). Endocrine System DevelopmentA number of studies evaluated effects on thyroid hormones in neonatal serum or cord blood associatedwith developmental exposure to PBDEs; however, findings are inconsistent between studies and congeners.Abdelouahab et al. (2013) evaluated the potential associations between thyroid hormone levels in theumbilical cord blood and maternal serum concentrations of PBDEs collected at <20 weeks of pregnancy(for ∑PBDE, median=30.92 ng/g lipid, maximum=726.09 ng/g lipid, n=380). Significant negative associations were observed between maternal PBDE levels and both free and total T4 in cord blood, butnot free or total T3 or TSH in cord blood. Similarly, neonatal TSH assessed in blood samples collected24 hours after birth (on average) was not related to PBDE concentrations (median ∑PBDE=25.4 ng/glipid) in maternal serum collected at the start of the third trimester or at delivery from 289 expectant mothers living in the Salinas Valley of California (Chevrier et al. 2011). In a low-exposure Korean birthcohort (n=104), no significant associations were observed between neonatal free or total T3, free or totalT4, or TSH and maternal PBDE concentrations in maternal serum collected at delivery (median∑PBDE=2.2 ng/g lipid) (Kim et al. 2015) PBDEs 2093. HEALTH EFFECTSOther studies of thyroid hormone changes in infants looked at hormone levels in umbilical cord orneonatal blood in relation to PBDE concentrations in maternal breast milk. Kim et al. (2011a) reported a significant negative correlation between breast milk PBDE (BDE 28 only, mean ≈0.1 ng/g lipid) and freeT4 in cord blood. Other studies that measured PBDEs in breast milk to assess exposure found no association with TSH in neonatal blood collected 3 days after delivery in a Norwegian population of239 mother-baby pairs (median ∑PBDE in breast milk=1.91 ng/g lipid) (Eggesbo et al. 2011), noassociation with T3, T4, free T4, or TSH in cord blood in a population of 149 Taiwanese mothers (median∑PBDE in breast milk=3.38 ng/g lipid from 42 mothers in Central Taiwan and 3.13 ng/g lipid from107 mothers in Southern Taiwan) (Shy et al. 2012), and no association with free T4, total T3, or TSH in infant serum at 3 weeks or 3 months in 186 Swedish mother-child pairs (median breast milkconcentrations of tetra-pentaBDE [BDE 47, BDE 99, BDE 100] of 2.3 ng/g lipid and BDE 153 of0.48 ng/g lipid) (Lignell et al. 2016).Inconsistent findings were also observed when thyroid hormones and PBDEs levels were evaluated ininfant serum and/or cord blood. For example, Mazdai et al. (2003) found no correlations between PBDEconcentrations (for ∑PBDE, range=14–460 ng/g lipid, median=39 ng/g lipid, n=12) and thyroid hormonelevels (free and total T4 and free and total T3) in umbilical cord blood. Similarly, there was no correlation between PBDEs and thyroid hormone levels in umbilical cord blood in another study of 21 South Korean mothers undergoing Cesarean section (Kim et al. 2012a). In this study, ∑PBDE in cord blood ranged from 2.28 to 30.94 ng/g lipid, with a median of 12.04 ng/g lipid. Both T3 and free T3 in cord blood weresignificantly inversely related to PBDE in cord blood (median ∑PBDE=3.49 ng/g lipid) in another studyof 54 Taiwanese births (Lin et al. 2011). T4, free T4, and TSH were unaffected in this study. Wan et al. (2010) found no significant relationship between the PBDE metabolite, 6-OH-BDE-47 (which ranged from <4 to 127 pg/g wet weight, with a median of 26 pg/g wet weight), and T4 in cord serum in 26 pregnant South Korean women. In an analysis of 289 births at Johns Hopkins Hospital in Maryland, PBDE in cord blood (median ∑PBDE=18.7 ng/g lipid) was compared to T4, free T4, and TSH in cordblood and T4 from blood spots collected from newborns at 2 and 18 days of age (on average) (Herbstman et al. 2008). There were consistent negative associations between PBDEs and free and total T4 in cord blood and/or spot samples, although these were primarily nonsignificant. Kim et al. (2011d, 2012b)analyzed blood samples collected from infants in neonatal screening tests. They found a positiverelationship between PBDEs with TSH (BDE 197 and BDE 196 only), a negative association with T3 (BDE 154 only) for control babies (for ∑PBDE, range=1.61–252.9 ng/g lipid, mean=56.70 ng/g lipid, n=12), and no significant relationships between PBDEs and thyroid hormones in babies with congenital PBDEs 2103. HEALTH EFFECTShypothyroidism (for ∑PBDE, range=2.22–861.0 ng/g lipid, mean=59.84 ng/g lipid, n=26). In anotherKorean birth cohort (n=104), significant positive associations were observed between neonatal serumTSH and cord BDE 99 levels and cord TSH and cord BDE 47 levels; no significant associations were observed for free or total T3 or T4 and ∑PBDE, BDE 47, or BDE 99 levels in cord blood (median ∑PBDE in cord=8.8 ng/g lipid) (Kim et al. 2015).Immune System DevelopmentA study of developmental immunity found reduced risk of atopic dermatitis in Japanese infants(diagnosed at 7 months of age by questionnaire) with higher concentrations of PBDE in umbilical cord blood (median ∑PBDE=41 pg/g wet weight for 27 cases and 54.5 pg/g wet weight for 54 controls)(Ochiai et al. 2014). Animal StudiesTeratology, Fetotoxicity, and Physical Growth and DevelopmentLower-brominated PBDEs: No exposure-related developmental changes were observed followinggestational exposure to pentaBDE in GD 12 rat embryos from dams exposed to doses up to 120 mg/kg/day via gavage from GD 6.5 to 11.5 (Ellis-Hutchings et al. 2009), in GD 20 rat fetuses fromdams exposed to doses up to 120 mg/kg/day via gavage from GD 6.5 to 19.5 (Ellis-Hutchings et al.2009), or in GD 20 rat fetuses from dams exposed to doses up to 200 mg/kg/day via gavage from GD 6 to 15 (Argus Research Laboratories 1985a). End points evaluated included fetal/embryo survival,resorptions, fetal weight and length, gross abnormalities, and skeletal and soft tissue abnormalities. Inthese studies, maternal toxicity (significantly decreased maternal weight gain) was observed at doses ≥100 mg/kg/day (Argus Research Laboratories 1985a; Ellis-Hutchings et al. 2009). Similarly, rat damsexposed to pentaBDE at doses up to 30.6 mg/kg/day from GD 6 to 21 did not show any exposure-relatedchanges in pregnancy or birth indices (Branchi et al. 2001, 2002, 2005; Kodavanti et al. 2010; Zhou et al. 2002). Zhou et al. (2002) also reported no change in offspring viability and growth, as assessed bynumbers of pups at birth and on PNDs 4–21, body weight of pups on PNDs 4–90, and eye opening statuson PNDs 11–18; however, Kodavanti et al. (2010) observed significantly decreased body weight fromPND 29 to 58 in female offspring at ≥10.2 mg/kg/day (8–10% reduction at PND 60). No exposure-related changes were observed in litter size, live births per litter, sex ratio, implantation sites, timing ofeye opening, or body or organ weights (brain, liver, thymus) on PND 21 in offspring from rat dams fed PBDEs 2113. HEALTH EFFECTSvanilla wafers dosed with commercial pentaBDE (DE-71) at doses up to 11.4 mg/kg/day from 28 dayspre-mating through PND 21 (Poon et al. 2011). In a similar study, no exposure-related changes were observed in litter size, live pups per litter, sex ratio, offspring survival, or timing of eye opening or pinnadetachment in offspring from rat dams fed cookies dosed with DE-71 at doses up to 30 mg/kg/day fromGD 1 to PND 21; however, a significant decrease in offspring body weight was observed during early development (up through PND 30) at 30 mg/kg/day, but not at doses ≤3 mg/kg/day (Bowers et al. 2015). Several studies examined developmental effects of very low doses of pentaBDE. Following maternalexposure to pentaBDE at doses of 0, 0.5, 1, or 2 mg/kg/day via gavage from GD 6 to 19, GD 20 ratfetuses showed delayed ossification and an increased incidence of internal variations at 2 mg/kg/day(Blanco et al. 2012). Significantly increased skeletal variations included delayed ossification of parietaland occipital bones, caudal vertebrae, and floating ribs. Soft tissue variations included significantincrease in the size of the ventricles of the heart and liver enlargement. No significant changes were observed in the number of live fetuses, the sex ratio, the average fetal body weight/litter, or externalmalformations (Blanco et al. 2012). In another study, incisor eruption was delayed in offspring from ratdams exposed on GD 6 to a single gavage dose of pentaBDE at 0.3 mg/kg, but not 0.06 mg/kg (averageage at eruption was not reported) (Kuriyama et al. 2005). No exposure-related delays were reported forfur development or eye opening (Kuriyama et al. 2005). Following single gavage doses of pentaBDE up to 0.3 mg/kg or tetraBDE up to 0.7 mg/kg/day in F0 rats on GD 6, no significant, exposure-relatedchanges were observed in F1 body weights through PND 100 or F2 fetal end points (implantation sites,implantation sites/dam, number of live fetuses, fetuses/dam, mean fetal weight, resorption rate, orincidence of skeletal anomalies) on GD 21 (Talsness et al. 2005, 2008).In several other studies assessing offspring body weight following developmental exposure to pentaBDE(without assessing any teratogenic or fetotoxic end points), no consistent body weight effects were found.No dose-related body weight changes were observed in offspring of male and female rats exposed topentaBDE at doses up to 25 mg/kg/day via gavage for 70 days pre-mating through PND 42 (Bondy et al. 2011, 2013). No exposure-related body weight effects were observed in rat pups from dams exposed to pentaBDE at doses up to 30 mg/kg/day via gavage or pentaBDE-dosed cookies from GD 6 to PND 18 or21 (Bansal et al. 2014; Cheng et al. 2009; Ellis-Hutchings et al. 2006; Miller et al. 2012; Zhao et al. 2014), or in mouse pups from dams exposed to pentaBDE at 452 mg/kg/day every third day from GD 4 toPND 17 via gavage (Skarman et al. 2005). A poorly-reported study indicated that pup weight wassignificantly decreased at PND 21 by 8, 16, and 15% following maternal exposure to pentaBDE at 50, 100, at 200 mg/kg/day, respectively, from GD 6 to PND 21 via gavage; by PND 63, no body weight PBDEs 2123. HEALTH EFFECTSeffects were observed (Hong et al. 2010). Another study reported significantly elevated pup body weighton PND 21 in rat offspring from dams exposed to pentaBDE at 2 mg/kg/day via gavage from GD 6 to PND 21; no changes were observed at 1 mg/kg/day (Blanco et al. 2014). Three studies exposed rat dams to different commercial octaBDE mixtures at 0, 2.5, 10, or 25 mg/kg/dayvia gavage on GDs 6–15, including FR-1208 (Life Science Research Israel Ltd. 1987), Saytex 111 (ArgusResearch Laboratories 1985b), and DE-79 (WIL Research Laboratories 1986). Following exposure to FR-1208, postimplantation loss was significantly elevated at 10 and 25 mg/kg/day when Freeman-Tukey arcsine transformed values were tested as normally distributing data (Student’s t-test); however, no statistical differences were observed with untransformed data and values were within the range oflaboratory historical control values. There were no indications of skeletal malformation or variations ordelayed or retarded ossification in any dose group (Life Science Research Israel Ltd. 1987). Followingexposure to Saytex 111, the number of resorptions per litter was significantly increased by 8-fold and thenumber of live fetuses per litter was decreased 6.6% in the 25 mg/kg/day group (Argus Research Laboratories 1985b). Average fetal body weights were also significantly reduced at 25 mg/kg/day (ArgusResearch Laboratories 1985b). Following exposure to DE-79, effects observed at 50 mg/kg/day included significantly reduced mean maternal body weight gain during the post-treatment period (GDs 16–20) and fetotoxicity as indicated by increased postimplantation loss due to late resorptions (not significantlyincreased compared to control group but exceeded historical control range), 39% reduced mean fetalweight (p<0.01), skeletal variations (e.g., reduced ossification of the skull and various unossified bones)that were associated with the reduced fetal weights in this group, and single instances of malformations (fetal anasarca, bent limb bones, unilateral absence of 13th rib) commonly associated with maternaltoxicity (WIL Research Laboratories 1986). In rabbit does exposed to octaBDE (Saytex 111) at 0, 2, 5, or 15 mg/kg/day on GDs 7–19 via gavage, the 15 mg/kg/day group showed slight fetotoxicity, as indicated by a significantly (p≤0.05) increased incidence of delayed ossification of the sternebrae (Breslin et al. 1989). This finding was accompanied byevidence of slight maternal toxicity as indicated by decreased body weight gain during GDs 7–20 and 7– 28 (not statistically identified), reduced body weight on GD 28 (7% less than controls, p≤0.05), and significantly increased absolute and relative liver weights on GD 28 (Breslin et al. 1989).Developmental effects were assessed in mice from dams fed cornflakes containing tetraBDE at doses of0, 0.03, 0.1, or 1 mg/kg/day from 28 days premating to PND 21 (Ta et al. 2011). At PND 21, asignificant decrease in crown-rump length of pups was observed in the 0.1 mg/kg/day group, compared PBDEs 2133. HEALTH EFFECTSwith controls; however, no other dose-group showed this effect and the ponderal index (weight/crownrump length) was not statistically different among treatment groups. Similarly, body weight wassignificantly decreased by ~13–17% from PND 14 to 18 in the 0.1 mg/kg/day group, compared with controls, but not in the 0.03 or 1 mg/kg/day groups. No changes were observed in gestation length, littersize, or sex ratio (Ta et al. 2011). In another study using the same protocol, no changes were observed in the number of pups per litter, the sex ratio, or pup body weight (Koenig et al. 2012). Pre-weaningweights were significantly reduced by ~10% in female offspring of female mice exposed to tetraBDE at0.03 mg/kg/day via gavage for 28 days premating through PND 21, but only in one of two experimentalreplicates (Woods et al. 2012). In a dietary study, no exposure-related changes were observed inpregnancy length, litter sizes, pup mortality, sex ratio, or pup weight in rats exposed to tetraBDE at dosesup to 32 mg/kg/day from GD 1 to PND 14 (Wang et al. 2011a).DecaBDE: Developmental effects were assessed in mice following maternal exposure to decaBDE atgavage doses of 0, 150, 750, 1,500, or 2,000 mg/kg/day from GD 7 to 9 (Chi et al. 2011). No grossexternal malformations were observed; however, significant exposure-related changes were observed inpostimplantation loss, resorptions, number of live litters, and fetal body weight on GD 16. Thepercentage of postimplantation loss per litter was significantly increased by 3, 2.7, and 9.8% at 750, 1,500, and 2,000 mg/kg/day, respectively, compared with control. At 1,500 and 2,000 mg/kg/day, thepercentage of resorptions per litter was also significantly increased by 2.7 and 8.6%, respectively.Additionally, the percentage of live fetuses per litter was significantly decreased by 10% in the high-dosegroup. Fetal body weight was significantly decreased by 10, 10, and 22% at 750, 1,500, and 2,000 mg/kg/day, respectively, compared with controls (Chi et al. 2011). A significant 15% reduction in pup body weight was observed in offspring of mouse dams fed dietary decaBDE at 260 mg/kg/day fromGD 10 through PND 21 (Watanabe et al. 2008); however, another study using the same protocol did notobserve body weight effects in PND 21 pups at doses up to 2,900 mg/kg/day (Watanabe et al. 2010b). Noexposure-related changes were observed in the number of liters or survival rate of pups in either study atdoses up to 3,100 mg/kg/day (Watanabe et al. 2008, 2010b). In other studies, no exposure-relatedteratogenic, fetotoxic, or body weight effects, and/or delays in attainment of developmental landmarks, were observed following decaBDE exposure in rat offspring from dams exposed to gavage doses up to 146 mg/kg/day from GD 10 to PND 20 or 21 (Fujimoto et al. 2011; Saegusa et al. 2012), in rat offspring from dams exposed to gavage doses up to 1,000 mg/kg/day from GD 0 to 19 or from GD 6 to PND 21(Hardy et al. 2001, 2002; Biesemeier et al. 2011), in mouse offspring from dams exposed to gavage doses up to 1,500 mg/kg/day from GD 0 to 17 (Tseng et al. 2008, 2013), or in mouse offspring from damsexposed to doses up to 20 mg/kg/day via micropipette from PND 2 to 15 (Rice et al. 2007). In a poorly PBDEs 2143. HEALTH EFFECTSreported study, the study authors reported that pup weight was significantly decreased in mouse offspring at PND 21 following maternal exposure to decaBDE at 500, 2,500, and 12,500 mg/kg/day via gavagefrom GD 6 to PND 21; however, the included graph and table do not support that statement (Hong et al. 2010).A lower purity commercial decaBDE product (77% decaBDE, 22% nonaBDE, 0.8% octaBDE) used in the 1970s was fetotoxic in rats at high, albeit not maternally toxic, dose levels. Developmental effects were investigated in GD 21 rat fetuses from dams exposed to decaBDE at doses of 10, 100, or1,000 mg/kg/day by gavage on GDs 6–15 (Dow Chemical Co. 1985; Norris et al. 1975a). The numbersof fetuses with subcutaneous edema and delayed ossification of normally developed skull bones were significantly increased at 1,000 mg/kg/day. Resorptions were significantly increased at ≥10 mg/kg/day,but the increases were not dose-related and rates in the high dose group were comparable to historicalcontrol values. As discussed in Section 3.2.2.5 (Reproductive Effects), a one-generation study of the 77% commercial decaBDE mixture at ≤100 mg/kg/day in rats found no effects on numbers of live pups at birth or during lactation, body weights of pups at birth or weaning, or skeletal development or soft-tissue histology of pups at weaning (Dow Chemical Co. 1975; Norris et al. 1975a).An additional study evaluated developmental effects in neonatal mice engineered to express neonatalhuman apolipoprotein E (three genotypes: apoE2, apoE3, and apoE4) exposed once to decaBDE at 0, 10, or 30 mg/kg via micropipette on PND 10 (Reverte et al. 2014). Investigators were assessing the potentialeffect of different human apoE genotypes on susceptibility to decaBDE exposure during development, asapolipoprotein is a genetic factor that is associated with varied vulnerability for the development ofneurodegenerative disease. No exposure-related changes were observed in survival, body weight gain, ontogeny of reflexes, pinna detachment, or incisor eruption. However, eye opening was significantlydelayed in apoE2 mice exposed to 30 mg/kg, compared with apoE2 controls.NeurodevelopmentLower-brominated PBDEsNeurobehavior: In a series of one-day exposure neonatal gavage studies using similar experimentaldesigns, mice exposed to pentaBDE at doses ≥0.8 mg/kg on PND 3 or PND 10 consistently showed alterations in open-field behavior at 2–8 months of age, characterized by decreased activity during thefirst 20-minute period followed by increased activity during the third 20-minute period; exposure-related PBDEs 2153. HEALTH EFFECTSeffects were not observed in any study at ≤0.4 mg/kg/day or following exposure to pentaBDE at 8 mg/kgon PND 19 (Eriksson et al. 2002b, 2006; Fischer et al. 2008; Hallgren et al. 2015; Sand et al. 2004;Viberg et al. 2002, 2004a, 2004b). These findings indicate an initial decrease in activity, but also a lackof habituation to new surroundings. The study authors noted that this nonhabituating behavior profile(i.e., decreased activity early in the test period and increased activity late in the test period) has also beenreported in adult mice neonatally exposed to certain ortho-PCB congeners (Eriksson and Fredriksson 1996a, 1996b). Several other 1-day exposure studies report similar findings in rats and mice followingexposure to various lower-brominated PBDEs. Decreased spontaneous activity and/or impaired habituation was observed in rats exposed to pentaBDE at 8 mg/kg on PND 10, mice exposed to hexaBDEat ≥0.45 mg/kg on PND 10, mice exposed to tetraBDE at 10.5 mg/kg on PND 10, mice exposed to heptaBDE at 15.2 mg/kg on PND 3, and mice exposed octaBDE at 16.8 mg/kg on PND 3 or 10 (Eriksson et al. 2001; Viberg et al. 2003a, 2005, 2006). Increased vertical activity was significantly increased at4 months, but not 2 months, in mice exposed to tetraBDE at ≥1 mg/kg on PND 10; no changes were observed in horizontal activity or habituation (Gee and Moser 2008). No changes in open-field behaviorwere observed in mice exposed to heptaBDE at 15.2 mg/kg or nonaBDE at 18.5 on PND 10 (Viberg et al. 2006). The observed effects may be modulated by the cholinergic system, as mice exposed to pentaBDEat 8 mg/kg on PND 10 showed significantly altered responses in a nicotine-induced behavior task(decreased instead of increased activity) (Viberg et al. 2002).Evidence for exposure-related changes in open-field behavior is less consistent in pre- and perinatalstudies. Male offspring from mouse dams exposed to pentaBDE at 18 mg/kg/day via gavage or “selfadministration” from a modified syringe from GD 6 to PND 21 showed significantly increased motoractivity during the 3rd 10-minute block of an open field test on PND 34, indicating decreased habituation (female offspring not evaluated) (Branchi et al. 2005). This effect was transient, as it was no longerobserved in male offspring at PND 60 to 120. No exposure related changes were observed in the amountof time spent in the center versus the middle of the open field (Branchi et al. 2005). However, male andfemale offspring from mouse dams exposed to pentaBDE at 0, 0.6, 6, or 30 mg/kg/day via gavage fromGD 6 to PND 21 showed changes suggestive of an age-dependent alteration in activity at ≥6 mg/kg/day;effects included hyperactivity (increased locomotion and rearing) and impaired habituation at PNDs 34 and 60, altered thigmotaxis (reduced time near walls) at PND 60, and a tendency to hypoactivity (reduced locomotion) at PND 120 (Branchi et al. 2001, 2002). In 24-hour observations of open-field behavior, total activity, time spent active, duration of activity per active phase, and total activity per active phase were all significantly increased in PND 36 male rat offspring following a single maternal exposure to0.3 mg/kg on GD 6 via gavage (Kuriyama et al. 2005). At PND 71, the increases in total activity and PBDEs 2163. HEALTH EFFECTStime spent active persisted in the 0.3 mg/kg group, and were also significantly increased at 0.06 mg/kg(Kuriyama et al. 2005). Bowers et al. (2015) reported a significant increase in rearing behavior in rats atPND 110 following maternal exposure to pentaBDE at doses up to 30 mg/kg/day via dosed cookie intake from GD 1 to PND 21; however, rearing behavior did not differ from control on PND 16, 55, or 230, and no other changes in motor activity or habituation behavior were observed in open-field testing at anytime-point. Kodavanti et al. (2010) also reported a lack of exposure-related changes in open-fieldbehavior in rats at PND 100, 114, or 273 following maternal exposure to pentaBDE at doses up to 30.6 mg/kg/day via gavage from GD 6 to PND 21. Similarly, no exposure-related changes were observedin general motor activity in an open field in PND 22 male and female offspring of rat dams exposed to pentaBDE at doses up to 2 mg/kg/day via gavage from GD 6 to PND 21; however, the offspring exposed to 2 mg/kg/day spent a significantly greater percentage of time in the center of the open field, indicatingdecreased anxiety (anxiolytic effect) (Blanco et al. 2013). However, no effect on anxiety behavior wasobserved in the emergence latency assay in PND 35 or 80 rat offspring of dams fed pentaBDE-dosed cookies at intake levels up to 30 mg/kg/day from GD 1 to PND 21 (Bowers et al. 2015). In offspring ofmouse dams fed cornflakes containing tetraBDE at doses up to 1 mg/kg/day from pre-mating day 28 through PND 21, no changes in locomotor activity were observed at PNW 8 or 17 (Koenig et al. 2012).In contrast, a second study using the same protocol reported significantly decreased locomotion and distance travelled in open-field testing of PND 60 females exposed to ≥0.1 mg/kg/day; at 0.03 mg/kg/day,PND 60 female mice showed decreased activity in the center of the open field (Ta et al. 2011). Noexposure-related changes in open-field behavior were observed at PND 60 in males or PND 42 in eithersex (Ta et al. 2011). Open field behavior was not significantly altered in PND 21 male offspring ofmouse dams administered tetraBDE at 0.2 mg/kg/day from GD 8 to PND 21, compared with control (Kimet al. 2015).In 1-day exposure neonatal gavage studies, significant impairments were also observed in learning and memory in the Morris water maze test in mice exposed to pentaBDE at 0.8 mg/kg on PND 10, miceexposed to hexaBDE at ≥0.9 mg/kg on PND 10, and mice exposed to octaBDE at 16.8 mg/kg on PND 10, or in rats exposed to tetraBDE at ≥1 mg/kg on PND 10 (Fischer et al. 2008; He et al. 2009, 2011; Viberget al. 2003a, 2006). Observed changes included increased latencies to find a hidden platform during a4-day training period and decreased distance travelled in the quadrant containing the hidden platform compared with controls; no changes were observed during a reversal phase on the 5th day (when theplatform was moved to a new quadrant). Similarly, impairments in learning and memory were observed in the radial arm maze in mice exposed to pentaBDE at 0.8 mg/kg on PND 10, as evidence by theincreased number of re-entries into maze arms from which the food pellet had already been eaten (Fischer PBDEs 2173. HEALTH EFFECTSet al. 2008). No exposure-related effects on learning or memory were observed in the Morris water maze in mice exposed to octaBDE at 16.8 mg/kg on PND 3 or nonaBDE at 18.5 mg/kg on PND 10 (Viberg etal. 2006). Based on the observed deficits, some of the studies evaluated the density of nicotiniccholinergic receptors in the hippocampus. Significant decreases in receptor density of 7–31% wereobserved in mice exposed to penta- or hexaBDE at ≥9 mg/kg on PND 10 (Viberg et al. 2003a, 2004b, 2005). Another study reported a significant 23% decrease in cortical density of nicotinic cholinergicreceptors in mice exposed to pentaBDE at 0.8 mg/kg on PND 10; hippocampal density was not altered atthis dose (Fischer et al. 2008). The observed effects may be modulated by the cholinergic system, as mice exposed to pentaBDE at 8 mg/kg on PND 10 showed significantly altered responses in a nicotine-induced behavior task (decreased instead of increased activity) (Viberg et al. 2002).Learning and memory impairments have also been reported following pre- and perinatal exposures. Following maternal exposure to pentaBDE at 0, 1, or 2 mg/kg/day via gavage from GD 6 to PND 21, anonsignificant trend toward increased time to locate the hidden platform in the Morris water maze was observed on days 2–4 in PND 22 rat offspring from both dose groups; however, the increase was onlysignificant on day 4 in the 2 mg/kg/day group (Blanco et al. 2013). In another study, significantlyincreased latency to find the escape platform was observed in the Morris water maze on training days 2–3 in PND 34 offspring of rat dams exposed to pentaBDE at 2 mg/kg/day from GD 6 to PND 21 (Cheng etal. 2009). In a low-dose study, maternal exposure to 0.2 mg/kg/day from GD 1 to PND 21 via gavage did not impair spatial learning in PND 34–36 offspring (Zhao et al. 2014). Effects observed at higher dosesmay be transient, as no exposure-related changes were observed in the Morris water maze in PND 235 ratoffspring following maternal exposure to pentaBDE at doses up to 30 mg/kg/day via dosed cookie intakefrom GD 1 to PND 21 (Bowers et al. 2015). No exposure-related deficits in the Morris water maze were observed during PNW 7 or 11 in offspring of mouse dams exposed to tetraBDE at 0.03 mg/kg/day viagavage from pre-mating day 28 to PND 21 (Woods et al. 2012). Similarly, offspring of mouse dams fed cornflakes containing tetraBDE at 0, 0.03, 0.1, or 1 mg/kg/day from pre-mating day 28 through PND 21 did not show exposure-related impairments in the Morris water maze at PNW 8 (Ta et al. 2011).However, another study using the same protocol reported a significant increase in the latency to find a hidden escape hole in the Barnes maze on the first day of training at ≥0.03 mg/kg/day in PNW 8 offspring (Koenig et al. 2012). No exposure-related effects were observed during trial days 2–4 or during reversallearning on day 5 (escape hole placed in a different location) at doses up to 1 mg/kg/day (Koenig et al.2012). Other studies assessed learning and attention using 5-choice visual learning and attention tasks conducted from PND 30 to 95 in offspring of rat dams exposed to pentaBDE via gavage from PND 6 to 12 (Driscoll et al. 2012; Dufault et al. 2005). Offspring exposed to 30 mg/kg/day required significantly PBDEs 2183. HEALTH EFFECTSmore trials to reach “passing” criterion in the visual discrimination task, and committed significantlymore errors than controls prior to reaching the criterion (Dufault et al. 2005). The number of omission errors (no choice made) was also significantly increased in rats exposed to 30 mg/kg/day. No exposure-related impairments were observed in the visual discrimination task at doses ≤15 mg/kg/day, and no exposure-related impairments were observed in the sustained attention task at doses up to 30 mg/kg/day(Driscoll et al. 2012; Dufault et al. 2005). In the three-chambered sociability task, general sociability (time spent in the chamber with a social target)was decreased in female offspring of mouse dams exposed to tetraBDE at 0.03 mg/kg/day via gavagefrom pre-mating day 28 to PND 21 (only dose evaluated); no exposure-related changes were observed inmale offspring (Woods et al. 2012). No changes were observed in social novelty (time spent in thechamber with a new social target vs. a familiar social target) or barrier social interaction tests (time spentinteracting with social target) in either sex (Woods et al. 2012). No exposure-related changes in socialinteraction were observed in offspring of mouse dams fed cornflakes containing tetraBDE at 0, 0.03, 0.1, or 1 mg/kg/day from pre-mating day 28 through PND 21 (Ta et al. 2011). No performance deficits wereobserved in additional neurobehavioral tests in these studies, including the elevated plus maze, acousticstartle and prepulse inhibition, and fear conditioning (Ta et al. 2011; Woods et al. 2012). Similarly, nosignificant, exposure-related changes were observed in social interaction or social novelty indices inPND 70 offspring of dams administered tetraBDE at 0 or 0.2 mg/kg/day via gavage from GD 8 to PND 21 (Kim et al. 2015). Additionally, no exposure-related effects were observed in social dominancetesting (tube test, urine marking test).Sensory and motor development. Sensory and motor development (righting reflex, forelimb stickgrasping reflex, forelimb placing reflexes, negative geotaxis, screen grasping and climbing, pole grasping, ultrasonic vocalizations, homing test) were assessed in male and female offspring of mouse dams exposed to pentaBDE at doses of 0, 0.6, 6, or 30 mg/kg/day via gavage from GD 6 to PND 21 (Branchi et al. 2001,2002). The screen climbing response was delayed by approximately 2 days in the 30 mg/kg/day group, compared with controls; no other exposure-related changes were observed in reflex development (Branchiet al. 2001, 2002). However, in rats, development of the cliff drop and negative geotaxis reflexes wassignificantly delayed in male pups following maternal exposure to pentaBDE at 2 mg/kg/day via gavagefrom GD 6 to PND 21; no delays were observed in the development of the righting reflex (Cheng et al. 2009). Similarly, development of the cliff drop reflex was significantly delayed in male rat pupsfollowing a single maternal exposure to pentaBDE at 0.3 mg/kg on GD 6, but not at 0.06 mg/kg (averageage at which the reflex developed was not reported) (Kuriyama et al. 2005). However, no exposure PBDEs 2193. HEALTH EFFECTSrelated changes were observed in the ability of male offspring to stay on a rotating rod for 3 minutes atdoses up to 0.3 mg/kg (Kuriyama et al. 2005). In a low-dose study, no exposure-related changes in reflex maturation or motor coordination were observed in PND 3–30 offspring of dams exposed to pentaBDE at0.2 mg/kg/day from GD 1 to PND 21 via gavage (Zhao et al. 2014). Similarly, no changes in grip strength or motor coordination (beam test) were observed in PND 12–60 rat offspring of dams fed pentaBDE-dosed cookies at intake levels up to 30 mg/kg/day from GD 1 to PND 21 (Bowers et al. 2015). However, the acoustic startle response was significantly increased at the highest intensity levels in rats onPND 9, but not PND 20 (Bowers et al. 2015). Sensory and motor development was also tested in offspring of mouse dams fed cornflakes containingtetraBDE at doses of 0, 0.03, 0.1, or 1 mg/kg/day from pre-mating day 28 to PND 21 (Ta et al. 2011). No significant exposure-related effects were found in the Wahlsten battery for sensory and motordevelopment (righting reflex, cliff aversion, needle grasp, visual placing, vibrissa placing, eye opening, ear opening, ear twitch response, screen pull, screen cling/climb, startle reflex) on PNDs 8–18 or rotarod performance on PNDs 35–36; however, significant differences in ultrasonic pup vocalization (UPV)measures were observed between the 1 mg/kg/day group and control, including increased duration ofUPVs on PND 13 and increased UPV bout duration on PND 9 and 13. A nonsignificant trend was also observed in the 1 mg/kg/day group in the number of UPVs on PNDs 13 and 17 (Ta et al. 2011). Similarly, female offspring of mouse dams exposed to tetraBDE at 0 or 0.03 mg/kg/day via gavage frompre-mating day 28 to PND 21 showed significantly decreased UPV between PND 8 and 16; no exposure-related deficits were observed in male or female offspring in sensory or motor test batteries from PND 8to 21 or rotarod performance at PNDs 29–30 or 66–68 (Woods et al. 2012). Koenig et al. (2012) also reported a lack of exposure-related changes motor tests (grip strength, ladder walk, gait analysis) inPND 14–42 offspring of mouse dams fed cornflakes containing tetraBDE at doses up to 1 mg/kg/dayfrom pre-mating day 28 through PND 21 (Koenig et al. 2012).In other studies, no exposure-related changes were observed in functional observation batteries in maleoffspring of rat dams exposed to pentaBDE at doses up to 30.6 mg/kg/day via gavage from GD 6 toPND 21 (assessed at PND 24 or 60) or male mice exposed to single tetraBDE doses up to 30 mg/kg viagavage on PND 10 (assessed at PNDs 12–18 or at 1 or 3 months of age) (Gee and Moser 2008; Kodavantiet al. 2010). Electrophysiology. Following a single exposure to tetraBDE at 0, 6.8, or 68 mg/kg via gavage on PND 10, hippocampal slices from PND 17 to 19 male mice were prepared for field-excitatory PBDEs 2203. HEALTH EFFECTSpostsynaptic potential (fEPSP) recordings (Dingemans et al. 2007). No exposure-related differences inthe stimulus-response relation were seen in fEPSPs. However, post-tetanus potential (PTP) and long-term potential (LTP) were significantly decreased in the hippocampus of exposed mice, compared with controls. No exposure-related effects on paired-pulse inhibition were observed.To assess auditory function, distortion product otoacoustic emissions (DPOAEs) were measured in adultoffspring of rat dams fed vanilla wafers containing pentaBDE at doses of 0, 5.7, or 11.4 mg/kg/day frompre-mating day 28 to PND 21 (Poon et al. 2011). No exposure-related changes were observed in DPOAEamplitudes, frequencies, or thresholds.Histology and organ weight. Ultrastructural changes were observed in the hippocampus of 2-month-old rats exposed once to tetraBDE on PND 10 at ≥5 mg/kg, but not at 1 mg/kg (He et al. 2009). At 5 mg/kg, the endoplasmic reticulum appeared increasing swollen and degranulated, and at 10 mg/kg, the neuronswere acutely affected, with puffed periplast, dissolved cell organelles, and vacuolized mitochondria (no incidence data reported) (He et al. 2009). No exposure-related changes in hippocampal histology wereobserved in PND 70 offspring of mouse dams fed cornflakes containing tetraBDE at 0, 0.03, 0.1, or1 mg/kg/day from pre-mating day 28 through PND 21 (Ta et al. 2011).No exposure-related changes in brain weight were observed in neonatal, weanling, or adult offspringfollowing maternal exposure to pentaBDE at 18 mg/kg/day via gavage on GD 6 to PND 21 (Ellis-Hutchings et al. 2006), maternal exposure to pentaBDE-dosed cookies at doses up to 30 mg/kg/day fromGD 1 to PND 21 (Bowers et al. 2015), or maternal exposure to pentaBDE-dosed vanilla wafers at dosesup to 11.4 mg/kg/day from pre-mating day 28 to PND 21 (Poon et al. 2011). Similarly, cerebrum weightwas unaffected in 2-month-old rats exposed once to tetraBDE at doses up to 10 mg/kg via gavage on PND 10 (He et al. 2011).Biochemical, proteomic, genomic, and epigenetic changes. Hippocampal and cortical tissues of rat pupswere evaluated for mRNA and protein expression levels of several T3-mediate proteins (BDNF, NCAM1, and GAP-43) following maternal exposure to tetraBDE at 0, 3.2, or 32 mg/kg/day via gavage from GD 1 to PND 14 (Wang et al. 2011a). At PNDs 1, 7, and 14, multiple alterations in mRNA and protein levelswere observed; however, magnitude, direction, and significance of the changes were time- and region-dependent, with no clear exposure-related pattern (Wang et al. 2011a). In another study, hippocampaland cortical tissues were evaluated for gene expression levels of TRα1, TRα2, TRβ1, and BDNF in PND 21 offspring of rat dams exposed to pentaBDE at 0, 1, or 2 mg/kg/day via gavage from GD 6 to PBDEs 2213. HEALTH EFFECTSPND 21 (Blanco et al. 2013). The only exposure-related change observed was a significant decrease inBDNF expression in the hippocampus at 2 mg/kg/day (Blanco et al. 2013). In 2-month-old rats exposed once to tetraBDE at 0, 1, 5, or 10 mg/kg on PND 10, various dose-relatedchanges were observed in the mRNA and protein expression levels of apoptotic proteins in thehippocampus, including induction of caspase3, caspase12, and cytochrome C (He et al. 2009). Costa etal. (2015) reported similar induction of caspase3 in mice sacrificed 24 hours following exposure to 20 mg/kg of tetraBDE on PND 10, compared with control. Markers of oxidative stress were also elevatedin exposed mice, including malondialdehyde (MDA), 8-isoprostane, and reactive protein carbonyls, and tetraBDE-mediated induction of caspase3 was significantly greater in glutathione-deficient knockout mice(Gclm -/-) than in wild-type mice (Costa et al. 2015). In another 1-day neonatal exposure study, brain levels of calcium/calmodulin-dependent protein kinase II (CaMKII), NMDA interacting proteins PSD-95 and SAP97, AMPA receptor subunit GluR1, and NMDA receptor subunits NR1, NR2A, and NR2B weremeasured in PNDs 17–19 mice exposed to tetraBDE at 0, 6.8, or 68 mg/kg on PND 10 (Dingemans et al. 2007). Protein expression levels of NR2B, GluR1, and the autophosphorylated-active form of CaMKIIwere significantly decreased in exposed mice compared with control. No exposure-related changes were observed for total CaMKII, NR1, NR2A, PSD-95, or SAP97 protein expression levels. Additionally, K+-evoked catecholamine release was evaluated from chromaffin cells isolated from mice in the 0 and68 mg/kg groups; no significant exposure-related effects were observed (Dingemans et al. 2007).Hippocampal, cortical, and cerebellar tissues were evaluated for markers of oxidative stress in PND 37male offspring of rat dams exposed to pentaBDE at 0 or 2 mg/kg/day via gavage from GD 6 to PND 21 (Cheng et al. 2009). In the hippocampus of offspring of exposed dams, the specific activities of SOD andGSH-Px were significantly decreased, lipid peroxidase, H2O2, and NO generation were significantlyincreased, and the number of free radicals was significantly increased, compared with controls. No significant changes were observed for these measures in the cerebral cortex or cerebellum, with the exception of significantly increased H2O2 generation in the cerebellum. No changes in brain GSHconcentrations were observed (Cheng et al. 2009). Increased markers of apoptosis and oxidative stresswere also observed in the hippocampus of offspring of rat dams exposed to the commercial pentaBDEmixture DE-71 at 30.6 mg/kg/day via gavage from GD 6 to PND 14 (Kodavanti et al. 2015). In a low-dose study, no exposure-related changes were observed in cortical, cerebellar, or hippocampal levels ofSOD, GSH-Px, GSH, H2O2, NO, or lipid peroxidation in PND 37 male offspring of dams exposed to0.2 mg/kg/day via gavage from GD 1 to PND 21, compared with controls (Zhao et al. 2014). PBDEs 2223. HEALTH EFFECTSDNA methylation and mRNA levels of DNA methyltransferase (Dmnt1) were measured in adultoffspring of mouse dams exposed to tetraBDE at 0 or 0.03 mg/kg/day via gavage from pre-mating day 28 to PND 21 (Woods et al. 2012). Global DNA methylation was significantly decreased in female, but notmale, offspring; no changes were observed in Dmct1 transcript levels.Cortical and hippocampal cholinergic gene transcription was evaluated in 2-month-old mice exposed to pentaBDE at 12 mg/kg via gavage on PND 10 (Hallgren et al. 2015). Significant transcriptional findingsincluded increased cortical nAchR-β2, AChR5, and nAChR-α4 in exposed mice, compared with controls;a near-significant increase in cortical AchE transcription was also observed. No exposure-relatedcholinergic gene transcription changes were observed in the hippocampus of 2-month-old mice. Additionally, no exposure-related cholinergic gene transcription changes were observed in the cortex orhippocampus of mice sacrificed 24 hours postexposure (i.e., PND 11). Cholinergic effects were alsoevaluated in offspring from a one-generation study in mink (Bull et al. 2007). F0 females were exposed to dietary pentaBDE at 0, 0.01, 0.05, or 0.25 mg/kg/day from pre-mating day 28 to PNW 6. No kits wereborn in the 0.25 mg/kg/day group; for the other groups, 6 kits/group were sacrificed at PNW 6 and 10 kits/group continued dietary exposure until PNW 27 and were sacrificed at PNW 45. In both PNW 6 and 45 kits, ChE was determined in blood plasma and cerebral cortex. In the cerebral cortex, ACh, mAChR,and nAChR binding were also measured. No exposure-related effects were observed in any cholinergic measures in 6-week-old kits or 27-week-old juveniles.DecaBDENeurobehavior. As seen in the lower-brominated PBDE neurobehavioral section, decreased spontaneousactivity and impaired habituation in open-field testing were observed at 2 and 4 months in mice exposed once to decaBDE on PND 3 at doses ≥2.22 mg/kg (Buratovic et al. 2014; Johansson et al. 2008; Viberg etal. 2003b, 2007). Also as observed with pentaBDE, decaBDE-exposed mice showed significantly alteredresponses in a cholinergic-induced behavior task (decreased instead of increased activity) at ≥5.76 mg/kg on PND 3 (Buratovic et al. 2014; Johansson et al. 2008). Following exposure on PND 10 or 19, however,no exposure-related changes in open-field behavior were observed in mice exposed to decaBDE at doses up to 20.1 mg/kg (Viberg et al. 2003b). In contrast, significantly increased locomotor activity during thefirst 1.5 hours of a 2-hour observation period was observed in PND 70 males following exposure to decaBDE at 20 mg/kg/day via micropipette from PND 2 to 15 (Rice et al. 2007). No changes in locomotor activity were observed at PND 70 in females or 1 year in either sex at doses up to 20 mg/kg/day (Rice et al. 2007). In a perinatal exposure study, no exposure-related changes in motor PBDEs 2233. HEALTH EFFECTSactivity were observed in neonatal, weanling, juvenile, or adult offspring of rat dams exposed to decaBDEat doses up to 1,000 mg/kg/day via gavage from GD 6 to PND 21 (Biesemeier et al. 2011).Operant training and visual discrimination were assessed in young adult (3-month-old) and aging(16-month-old) mice following exposure to decaBDE at 0, 6, or 20 mg/kg/day via micropipette on PNDs 2–15 (Rice et al. 2009). Neurobehavioral tasks included lever-press training followed by a series of operant procedures, including a fixed-ratio (FR) schedule of reinforcement, a fixed-interval (FI) 2-minute schedule, and a light-dark visual discrimination. No significant, exposure-related effects were observed in the young adult cohort; however, learning impairment and impulsivity were observed in theaging cohort. In the FR task, exposure did not lead to impaired performance; however, there was asignificant main effect of decaBDE exposure on the slope parameter for the number of earned food pellets. The control group earned the fewest reinforcers at the beginning of the task, and the high-dosegroup earned the most; by the end of the 10 sessions, all dose groups earned about the same number ofreinforcers. In the FI task, there were significant main effects of decaBDE exposure on the fitted meanfor the overall response rate, with the high-dose having a marginally higher response rate than the controlgroup (p=0.06). This means that the exposed mice were emitting more responses for the same number ofreinforcers. No significant exposure-related changes were observed for other parameters measured in the FI task (pause time, run rate, index of curvature, or number of responses during the feed cycle). Anumber of significantly altered parameters were observed during the visual discrimination task (41 trials). Significantly decreased number of first choice errors and shorter response latencies were observed in6 and 20 mg/kg/day females and 20 mg/kg/day males; these findings were particularly pronounced in the earlier trials. However, mice in the 20 mg/kg/day group showed a significantly higher rate of error in thelast 15 trials, compared with controls. Additionally, after an initial error, mice in the 20 mg/kg/day group made significantly more “perseverative” errors than controls.Altered spatial learning and memory have been reported in the Morris water maze following prenatal orneonatal exposure to decaBDE. While initial spatial learning was not impaired in mice exposed to 5.76 or13.4 mg/kg on PND 3, reversal learning (ability to find the escape platform in a new location after initial training) was significantly impaired at ≥5.76 mg/kg at 5 and 7 months (Buratovic et al. 2014). Following gestational exposure from GD 1 to 14, rat offspring showed significantly impaired spatial learning on PND 25 at ≥30 mg/kg/day, but not at 10 mg/kg/day, compared with controls; reversal learning was notassessed in this study (Chen et al. 2014). An additional study evaluated spatial learning and memory in neonatal mice engineered to express neonatal human apolipoprotein E (three genotypes: apoE2, apoE3, and apoE4) exposed once to decaBDE at 0, 10, or 30 mg/kg via micropipette on PND 10 (Reverte et al. PBDEs 2243. HEALTH EFFECTS2013). Investigators were assessing the potential effect of different human apoE genotypes on susceptibility to decaBDE exposure during development, as apolipoprotein is a genetic factor that isassociated with varied vulnerability for the development of neurodegenerative disease. Impaired learningand memory were observed at 4 months in apoE3 and apoE4 male mice at ≥10 mg/kg/day and at12 months in apoE3 females exposed to 30 mg/kg (lack of preference for target quadrant). No wild-typemice were evaluated.In other behavioral tests, no exposure-related changes were observed in auditory startle responses, orlearning and memory in the water-filled T-maze in weanling, juvenile, or adult offspring from rat damsexposed to decaBDE at doses up to 1,000 mg/kg/day via gavage from GD 6 to PND 21 (Biesemeier et al. 2011) or in the elevated plus maze in 4-month-old mice exposed once to decaBDE at doses up to20.1 mg/kg via gavage on PND 3 (Johansson et al. 2008). No changes in were observed in surface-righting or geotaxis reflexes or motor function assays (tail pull, cling and climb test) at PNDs 12–16 in mice engineered to express neonatal human apolipoprotein E (three genotypes: apoE2, apoE3, and apoE4) exposed once to decaBDE at 0, 10, or 30 mg/kg via micropipette on PND 10 (Reverte et al. 2014).Sensory and motor development. A comprehensive functional observation battery was conducted everyother day from PND 2 to 20 in mice exposed to decaBDE at 0, 6, or 20 mg/kg/day via micropipette fromPND 2 to 15 (Rice et al. 2007). The only exposure-related changes observed were a significant reductionin the number of male and female pups performing the palpebral reflex on PND 14 and a significantreduction in the number of male pups performing an effective forelimb grip on PNDs 14 and 16 in the20 mg/kg/day group.Electrophysiology. Five groups of developing mice were exposed to decaBDE at 0 or 20.1 mg/kg/day viagavage during the following periods: gestational (GDs 1–21, via dam), lactational (PNDs 1–21, via dam), neonatal (PNDs 3–21, direct), post-weaning (PNDs 22–41), or gestational, lactational, and post-weaning(GD 1–PND 21, via dam, and PNDs 22–41, direct) (Xing et al. 2009). In the control group, dams wereadministered the vehicle only from GD 1 to PND 21 and offspring were administered the vehicle fromPND 22 to 41. The dams and pups in all exposure groups were administered the vehicle during non-exposure periods (e.g., the group exposed to decaBDE during gestation only was administered the vehicledaily from PND 1 to 41). In all groups, in vivo extracellular recording of synaptic transmission in the hippocampus was measured on PND 60 by placing the anesthetized mice in a stereotaxic head holder. Significantly decreased synaptic potency, short-term plasticity, and long-term potentiation were observed in exposed mice from the neonatal, post-weaning, and gestation+lactation+post-weaning groups. In the PBDEs 2253. HEALTH EFFECTSlactation-only group, long-term potentiation was significantly decreased, but no exposure-related changes were observed in synaptic potency or short-term plasticity. No exposure-related changes were observedin the gestation-only exposure group.Histology and organ weight. No exposure-related changes were observed in brain weight, histology, ormorphometry in weanling or juvenile offspring from rat dams exposed to decaBDE at doses up to1,000 mg/kg/day via gavage from GD 6 to PND 21 or PND 10 to 21 (Biesemeier et al. 2011; Fujimoto et al. 2011; Saegusa et al. 2012) or juvenile offspring from rat dams exposed to decaBDE at doses up to 1,500 mg/kg/day via gavage from GD 0 to 17 (Tseng et al. 2008). Absolute brain weight wassignificantly decreased in GD 16 fetuses following maternal exposure to decaBDE at 2,000 mg/kg/day viagavage on GDs 7–9; however, the difference was no longer significant once brain weights were adjustedfor body weight (Chi et al. 2011).Biochemical, proteomic, and genomic changes. Seven months following exposure to decaBDE at 0 or13.4 mg/kg on PND 3, the cortical and hippocampal expression of CaMKII, Gap-43, Tau, and synaptophysin were evaluated in male and female mice (Buratovic et al. 2014). Significant increases in protein levels of CaMKII, Gap-43, and Tau were observed in the cortex and hippocampus in exposed male mice and increased levels of Tau were observed in the cortex and hippocampus of exposed female mice, compared with control. No changes in synaptophysin were observed.Neonatal mice engineered to express neonatal human apolipoprotein E (three genotypes: apoE2, apoE3, apoE4) were exposed once to decaBDE at 0, 10, or 30 mg/kg via micropipette on PND 10 and evaluated for BDNF levels in the hippocampus (Reverte et al. 2013). When all genotypes were combined foranalysis, hippocampal BDNF levels were significantly elevated in males and females (combined) in the 30 mg/kg group. No wild-type mice were evaluated.Following exposure to decaBDE at 0, 2, 15, or 146 mg/kg/day from GD 10 to PND 21, brains fromoffspring sacrificed on PND 20 and 77 were fixed for hippocampal immunohistochemistry (reelin, flutamic acid decarboxylase 67 [GAD67], EphA5, Tacr3, and neuron-specific nuclear protein [NeuN])and cresyl violet staining of apoptotic bodies (Saegusa et al. 2012). Significant, exposure-related findings included an increase in the number of reelin-immunoreactive cells in the dentate hilus in groups exposedto ≥15 mg/kg/day on PND 20, an increase in the number of EphA5-postive cells in the CA1 layer onPND 20, and a minimal increase in the number NeuN-immunoreactive cells in the hilus in the146 mg/kg/day group on PND 77. PBDEs 2263. HEALTH EFFECTSReproductive System DevelopmentLower-brominated PBDEs: Reproductive performance was assessed in F1 male and female offspringfrom rat dams exposed to single pentaBDE doses of 0, 0.06, or 0.3 mg/kg or tetraBDE doses of 0, 0.14, or0.7 mg/kg on GD 6 via gavage (Kuriyama et al. 2005; Talsness et al. 2005, 2008). Following pentaBDEexposure, no significant exposure-related effects were observed in F1 male fertility or F2 litter parameters when exposed males were mated with unexposed females (Kuriyama et al. 2005). Additionally,ejaculatory and mounting latencies, intromission frequency and latency, and number of penetrations were not altered in F1 exposed mice, compared with controls; however, the percent of males with two or moreejaculations was significantly decreased by 32% in the F1 males from the 0.3 mg/kg group (Kuriyama etal. 2005). In pentaBDE-exposed F1 females mated to unexposed males, there were no exposure-relatedchanges in female pregnancy rate, total implantation sites, implantation sites/dam, F2 fetuses/gravid dam,or total number of live F2 fetuses (Talsness et al. 2005). However, the resorption rate was 12 and 15% in the 0.06 and 0.03 mg/kg groups, respectively, compared with the control rate of 9%, and the percentage oflitters with resorptions was 69 and 72% in the 0.06 and 0.3 mg/kg groups, respectively, compared with the control percentage of 47% (Talsness et al. 2005). Following tetraBDE exposure, reproductiveperformance of F1 females was unaltered by developmental tetraBDE exposure (Talsness et al. 2008).No exposure-related changes were observed for the following F2 litter parameters: total number ofimplantation sites, implantation sites/dam, number of live fetuses, fetuses/dam, mean fetal weight, orresorption rate. In the 0.7 mg/kg group, the sex ratio was significantly altered; however, comparison ofthe altered sex ratio with controls from two different historical experiments (n=24 and 43 litters) revealedno differences (Talsness et al. 2008). F1 male reproductive performance was not assessed followingtetraBDE exposure.Male offspring of rat dams exposed to pentaBDE at 0, 1.7, 10.2, or 30.6 mg/kg/day via gavage from GD 6 to PND 21 showed a significant 1.8-day delay in PPS (an external sign of pubertal development), compared with controls (Kodavanti et al. 2010). A 5.5% decrease in AGD was also observed at PND 7 in the 30.6 mg/kg/day group; although not statistically significant, the study authors argue that this findingmay be biologically relevant, as the findings were not confounded by body weight effects and wereaccompanied by the delay in PPS and a 20% decrease in mean testosterone concentration on PND 60 (Kodavanti et al. 2010). After maternal exposure to a lower pentaBDE dose (2 mg/kg/day) from GD 6 to PND 21 via gavage, no exposure-related changes in AGD were observed in male offspring on PND 1(Cheng et al. 2009). Female offspring of rat dams exposed to pentaBDE at 0, 1.7, 10.2, or PBDEs 2273. HEALTH EFFECTS30.6 mg/kg/day via gavage from GD 6 to PND 21 showed a significant reduction in mammary gland development on PND 21 at 10.2 and 30.6 mg/kg/day, showing lack of outgrowth, fewer lateral branchesand limited terminal end bud development, compared with controls (Kodavanti et al. 2010). Vaginalopening was not assessed. As discussed in Section 3.2.2.5 (Reproductive Effects), delays in reproductive development also occurred in male and female rats that were exposed to pentaBDE at 0, 3, 30, or 60 mg/kg/day via gavage duringpubertal development (PNDs 23–53 in males or PNDs 22–41 in females) (Stoker et al. 2004). PPS wassignificantly delayed by 1.7 and 2.1 days in the 30 and 60 mg/kg/day groups, respectively, and vaginalopening was significantly delayed by 1.8 days in the 60 mg/kg/day group. However, no changes wereobserved in estrous cycling at doses up to 60 mg/kg/day (Stoker et al. 2004).Following a single gavage administration of pentaBDE at 0, 0.06 or 0.3 mg/kg on GD 6 to rat dams, various exposure-related sperm effects were observed in male offspring on PND 140 in both dose groups(Kuriyama et al. 2005). Spermatid number and daily sperm production were significantly decreased by 31 and 34%, and sperm number was significantly decreased by 29 and 18%, at 0.06 and 0.3 mg/kg, respectively (Kuriyama et al. 2005).Following a single gavage administration of pentaBDE at 0, 0.06 or 0.3 mg/kg on GD 6 to rat dams, multiple ultrastructural changes were noted in the ovaries of female offspring at PND 90 in both exposuregroups, including destruction of the surface of the serosal epithelial cells, necrosis, and numerousvesicular structures with dense granular material within the cytoplasm (Talsness et al. 2005). Additionalchanges observed in the 0.3 mg/kg group included degenerative changes and aggregates of small andlarge vesicles filled with homogeneously dense granular material in the cytoplasm and clumpedchromatin within the condensed nucleus. No statistically significant, exposure-related histological changes were observed at the light microscopic level in the ovary, uterus, or vagina of female offspring,and no exposure-related effects were observed on the number of ovarian follicles (Talsness et al. 2005).Following a single gavage administration of tetraBDE at 0, 0.14, or 0.7 mg/kg on GD 6 to rat dams, themean number of secondary follicles in the ovaries of female offspring was significantly decreased by 43%in both exposure groups at PND 38, and the mean number of tertiary follicles was significantly decreasedby 38% in the 0.7 mg/kg group, compared with controls (Talsness et al. 2008). No exposure-relatedchanges were observed in the number of primordial, primary, or atretic follicles. No histopathologicallesions or organ weight changes were observed in the ovary, uterus, or vagina at PND 38 or 100 (Talsnesset al. 2008). However, ultrastructural changes of the ovaries were observed on PND 200 in F1 females PBDEs 2283. HEALTH EFFECTSfrom both exposure groups, including an accumulation of vesicular structures with homogeneously densegranular material in the cytoplasm of the stromal cells, which appeared to fuse together to form large vacuoles (Talsness et al. 2008). No exposure-related histopathological changes were observed in the uterior ovaries from female rats exposed to pentaBDE at gavage doses up to 60 mg/kg/day for 20 days duringpubertal development (PNDs 22–41) (Stoker et al. 2004).In 2-month-old females exposed to tetraBDE on PND 10, relative uterine weights were significantlydecreased by 23–36% at doses ≥1 mg/kg and relative ovary weights were significantly increased by 27– 35% at doses ≥5 mg/kg (He et al. 2011). In other studies, no exposure-related changes were observed in ovary or uterus weights in PND 100 offspring of rat dams exposed to tetraBDE at doses up to 0.7 mg/kgvia gavage on GD 6 (Talsness et al. 2008), in PND 31 offspring of rat dams exposed to pentaBDE at18 mg/kg/day from GD 6 to PND 18 via gavage (Ellis-Hutchings et al. 2006), in PND 43 offspring of ratdams exposed to pentaBDE at doses up to 25 mg/kg/day from pre-mating day 70 to PND 42 (Bondy et al. 2013), or in PND 42 rats exposed to pentaBDE at doses up to 60 mg/kg/day for 20 days during pubertaldevelopment (PNDs 22–41) (Stoker et al. 2004).No exposure-related histopathological changes were observed in the epididymides or left testis from male rats exposed to pentaBDE at gavage doses up to 60 mg/kg/day for 31 days during pubertal development(PNDs 23–53) (Stoker et al. 2004). No other developmental studies examined male reproductivehistology following exposure to lower-brominated PBDEs.Relative testes weights were significantly increased by 1.3-fold in PND 31 male offspring of rat dams exposed to pentaBDE at 18 mg/kg/day from GD 6 to PND 18 via gavage (absolute weights not reported);no exposure-related changes were observed on PND 12 and 18 (Ellis-Hutchings et al. 2006). In contrast, following a single maternal exposure to pentaBDE on GD 6, adult male rat offspring showed significant10 and 11% decreases in relative testes weights at 0.06 and 0.3 mg/kg, respectively, as well as a significant 5% decrease in relative epididymis weight at 0.3 mg/kg (Kuriyama et al. 2005). Neitherabsolute organ weights nor body weight were significantly altered (Kuriyama et al. 2005). In otherstudies, no exposure-related weight effects were observed in male reproductive organs (seminal vesicles,prostate, epididymides, or testes) in PND 60 offspring of rat dams exposed to pentaBDE at doses up to30.6 mg/kg/day from GD 5 to PND 21 (Kodavanti et al. 2010), in 2-month-old rats exposed to tetraBDEat doses up to 10 mg/kg on PND 10 (He et al. 2011), in PND 43 offspring of rat dams exposed topentaBDE from pre-mating day 70 to PND 42 at doses up to 25 mg/kg/day (Bondy et al. 2013), or in PBDEs 2293. HEALTH EFFECTSPND 53 rats exposed to pentaBDE at doses up to 60 mg/kg/day for 31 days during pubertal development(PNDs 23–53) (Stoker et al. 2004). After a single maternal gavage exposure to tetraBDE on GD 6, serum E2 levels in female rat offspringwere significantly decreased on PND 28 by ~38% at 0.7 mg/kg; no change was observed in the0.14 mg/kg group (Talsness et al. 2008). No exposure-related changes in serum testosterone or LH were observed in adult F1 males after a single exposure to pentaBDE at doses up to 0.3 mg/kg via gavage on GD 6 (Kuriyama et al. 2005). Exposure-related effects on serum testosterone levels were also notobserved in male offspring of rat dams on PND 60 following exposure to pentaBDE at doses up to30.6 mg/kg/day from GD 6 to PND 21 (Kodavanti et al. 2010). In male rats exposed to pentaBDE at 0, 3, 30, or 60 mg/kg/day via gavage during pubertal development (PNDs 23–53), serum prolactin wasincreased 2-fold in the 60 mg/kg/day group approximately 2 hours after the final exposure on PND 53;however, no exposure-related changes were observed in serum testosterone, serum or pituitaryluteinizing, or pituitary prolactin levels (Stoker et al. 2004). DecaBDE: Male offspring of mouse dams exposed to decaBDE at 0, 10, 500, or 1,500 mg/kg/day viagavage from GD 0 to 17 were assessed for reproductive system effects on PND 71 (Tseng et al. 2013).The mean AGD and AGI (anogenital index; corrected for body weight) were significantly reduced in the1,500 mg/kg/day group, compared with control. No exposure-related changes were observed in testicularindex ([testicular length x testicular width]/body weight) or male reproductive organ weights (testis,epididymis, cauda epididymis, or seminal vesicles). However, increased incidences of testicular lesions were observed in male offspring, with increased incidence of slight/moderate vacuolization in interstitialcells in all treated groups (0/5, 4/5, 3/5, or 5/5 at 0, 10, 500, or 1,500 mg/kg/day, respectively) andincreased incidence of slight-severe vacuolization in seminiferous tubules at 1500 mg/kg/day (4/5; controlincidence 0/5). Additionally, in the 1,500 mg/kg/day group, seminiferous tubules had lost almost allspermatozoa and spermatids. Analysis of male offspring sperm parameters (from the seminiferous tubules) showed a significant increase in percentage of abnormal sperm heads in the 1,500 mg/kg/daygroup (18.2%) compared with controls (10.3%). No changes were observed in abnormal sperm heads atlower doses, and no changes were observed in sperm count, motility, or velocity at any dose. However, evidence of sperm damage was observed in all exposed groups. The DNA fragmentation index (DFI) and level of sperm with DNA damage (X αT) were significantly elevated in all exposed groups in a dose-related manner. Sperm H2O2 generation was significantly elevated in a dose-related manner; however,only the values in the 10 and 1,500 mg/kg/day groups reached statistical significance in pairwise PBDEs 2303. HEALTH EFFECTScomparisons with controls. No changes were observed in sperm O2- generation. No changes in serumtestosterone levels were observed.In other studies, no exposure-related changes were observed in AGD, onset of puberty (PPS or vaginalopening), estrous parameters, and/or reproductive organ weight and histology in offspring of rat dams exposed to dietary decaBDE at doses up to 146 mg/kg/day from GD 10 to PND 21 (Fujimoto et al. 2011)or gavage doses up to 1,000 mg/kg/day from GD 6 to PND 21 (Biesemeier et al. 2011). Similarly, no exposure-related changes were observed in AGD or the onset of puberty (vaginal opening or descent ofthe testes) in mice exposed to decaBDE at doses up to 20 mg/kg/day via micropipette from PND 2 to 15 (Rice et al. 2007). No changes in were observed in vaginal opening or testicle descent in mice engineeredto express neonatal human apolipoprotein E (three genotypes: apoE2, apoE3, and apoE4) exposed onceto decaBDE at 0, 10, or 30 mg/kg via micropipette on PND 10 (Reverte et al. 2014).Endocrine System DevelopmentLower-brominated PBDEs: Histological, ultrastructural, and morphometric changes of the thyroid wereobserved at PND 100 in female offspring of dams exposed once to tetraBDE via gavage on GD 6 at 0, 0.14, or 0.7 mg/kg; however, a clear dose-response pattern cannot be determined from the available data (Talsness et al. 2008). The study authors reported “occasional” follicular cyst formation in the 0.14 mg/kg group with multiple areas of degenerated follicular epithelium. In the 0.7 mg/kg group, onlymild cyst formation was observed. Ultrastructural changes observed in both exposure groups included irregular, non-typical follicular shape and detached and swollen follicular cells. Incidence data were notreported for any of these end points. Morphometric analysis showed that the colloid area of the thyroid was also significantly increased in offspring, but only in the 0.14 mg/kg group (Talsness et al. 2008). In mink, no exposure-related changes were observed in thyroid histology in PNW 6 offspring of sowsexposed to pentaBDE at doses up to 0.06 mg/kg/day from pre-mating week 4 to PNW 6 (Zhang et al. 2009).In rats exposed once to tetraBDE via gavage on PND 10 at doses of 0, 1, 5, or 10 mg/kg, relative thyroid weights were significantly decreased by 11% in the 10 mg/kg group at 2 months of age (He et al. 2011). In other studies, no exposure-related changes in thyroid weight were observed in rat offspring of damsexposed once to tetraBDE at gavage doses up to 0.7 mg/kg on GD 6 (Talsness et al. 2008) or in minkoffspring of sows exposed to pentaBDE at doses up to 0.06 mg/kg/day from pre-mating week 4 to PNW 6(Zhang et al. 2009). PBDEs 2313. HEALTH EFFECTSSignificant reductions in serum T4 ranging from 12 to 90% were observed in male and female offspring ofrat dams exposed once to pentaBDE on GD 6 at 0.3 mg/kg (Kuriyama et al. 2007), exposed to pentaBDEor commercial pentaBDE mixtures from GD 6 to PND 18 or 21 via gavage at doses as low as2 mg/kg/day (Blanco et al. 2013; Ellis-Hutchings et al. 2006; Kodavanti et al. 2010; Miller et al. 2012;Shah et al. 2011; Szabo et al. 2009; Zhou et al. 2002), exposed to pentaBDE from GD 1 or 6 to PND 21 via dosed cookies at intake levels ≥3 mg/kg/day (Bansal et al. 2014; Bowers et al. 2015), or exposed todietary tetraBDE from GD 1 to PND 14 at doses ≥3.2 mg/kg/day (Wang et al. 2011a). These reductionswere observed in offspring between PND 7 and 22. In studies described above that also evaluatedoffspring at older ages, it was found that the effects were no longer observed at PNDs 31–60 (Bowers etal. 2015; Ellis-Hutchings et al. 2006; Kodavanti et al. 2010; Szabo et al. 2009; Zhou et al. 2002), indicating that serum T4 changes in offspring may be transient. In one-generation studies, offspring fromrat dams fed pentaBDE-dosed vanilla wafers from pre-mating day 28 to PND 21 showed significant 43– 55% reductions in serum T4 on PND 21 at ≥5.7 mg/kg/day (Poon et al. 2011) and male and female offspring from rat dams exposed to pentaBDE via gavage from pre-mating day 70 to PND 42 showed significant 65–70% reductions in serum T4 on PND 21 at 25 mg/kg/day (but not ≤5 mg/kg/day) (Bondy etal. 2011, 2013). In rats and mice exposed once to tetraBDE via gavage on PND 10 at doses up to 10 or20 mg/kg, respectively, no exposure-related changes were observed in serum T4 levels (Costa et al. 2015;Gee et al. 2008; He et al. 2011). In other mouse studies, maternal exposure to the pure pentaBDEcongener BDE 99 at doses up to 452 mg/kg/day via gavage from GD 6 to PND 21 or from GD 4 toPND 17 did not lead to altered serum T4 levels in offspring at PNDs 11–37 (Branchi et al. 2005; Skarman et al. 2005). In contrast, maternal exposure to the commercial pentaBDE mixture Bromkal 70-5DE at452 mg/kg/day via gavage from GD 4 to PND 17 caused a significant 29–32% reduction in serum T4 in PND 11 mouse offspring (Skarman et al. 2005). This effect was transient, as it was no longer observed atPND 18 (Skarman et al. 2005). In mink, no exposure-related changes were observed in serum T4 levels inPNW 6 offspring of sows exposed to dietary pentaBDE at doses up to 0.06 mg/kg/day from pre-matingweek 4 to PNW 6 (Zhang et al. 2009). In mink offspring that continued exposure through PNW 33, serum T4 levels were significantly increased by 31% compared with control; however, this response was entirely attributable to the females, which were significantly elevated by 71% compared to males at thesame dose (change compared with female controls was not reported) (Zhang et al. 2009).Evidence for exposure-related serum T3 changes is less consistent. In a one-generation study, male and female offspring from rat dams exposed to pentaBDE via gavage from pre-mating day 70 to PND 42showed significant 16–27% reductions in serum T3 on PND 21 at 25 mg/kg/day (but not ≤5 mg/kg/day) PBDEs 2323. HEALTH EFFECTS(Bondy et al. 2013). Similarly, offspring of rat dams exposed to pentaBDE at ≥2 mg/kg/day via gavagefrom GD 6 to PND 21 showed a significant 19–25% reduction in serum T3 levels on PNDs 21–23 and offspring from dams fed cookies dosed with ≥3 mg/kg/day pentaBDE from GD 1 to PND 21 showed asignificant 5–40% reduction in serum T3 levels on PND 21; no exposure-related effect was observed atdoses ≤1.7 mg/kg/day (Blanco et al. 2013; Bowers et al. 2015; Shah et al. 2011). This effect was transient, as it was no longer observed at PND 50 or later (Bowers et al. 2015; Shah et al. 2011). Incontrast, exposure-related changes in serum T3 levels were not observed in offspring of rat dams exposed to pentaBDE at doses up to 30.6 mg/kg/day via gavage or pentaBDE-dosed cookies from GD 6 to PND 18 or 21 (Bansal et al. 2014; Ellis-Hutchings et al. 2006; Kodavanti et al. 2010; Szabo et al. 2009) oroffspring of rat dams exposed to a single pentaBDE dose up to 0.3 mg/kg on GD 6 via gavage (Kuriyamaet al. 2007). In rats and mice exposed once to tetraBDE via gavage on PND 10 at doses up to 10 or20 mg/kg, respectively, no exposure-related changes were observed in serum T3 levels (Costa et al. 2015;Gee et al. 2008; He et al. 2011). In mink, no exposure-related changes were observed in serum T3 levels in PNW 6 offspring of sows exposed to pentaBDE at doses up to 0.06 mg/kg/day from pre-mating week 4to PNW 6; however, in offspring that continued exposure through PNW 33, serum T3 levels were significantly reduced by ~31% at 0.06 mg/kg/day (Zhang et al. 2009).No exposure-related changes in serum TSH were observed in offspring of rat dams exposed to a single pentaBDE dose up to 0.3 mg/kg on GD 6 via gavage (Kuriyama et al. 2005), offspring of rat damsexposed to pentaBDE at doses up to 30.6 mg/kg/day via gavage or pentaBDE-dosed cookies from GD 6to PND 21 (Bansal et al. 2014; Kodavanti et al. 2010), or rats exposed once to tetraBDE via gavage on PND 10 at doses up to 10 mg/kg (He et al. 2011).DecaBDE: Following maternal exposure to decaBDE at 0, 2, 15, or 146 mg/kg/day via gavage fromGD 6 to PND 21, the incidences of diffuse follicular cell hypertrophy in the thyroid of rat offspring on PND 21 were 0/10, 1/10, 3/10, and 9/10 in males and 0/10, 3/10, 2/10, and 4/10 in females, respectively(Fujimoto et al. 2011). No change was observed in thyroid weight (Fujimoto et al. 2011). Followingmaternal exposure to decaBDE at 0, 10, 500, or 1,500 mg/kg/day via gavage on GDs 0–17, histologicalevaluation of the thyroid glands in male offspring at PND 71 showed that a few acini were slightlyenlarged in the 1,500 mg/kg/day group, compared to the controls (female offspring were not evaluated)(Tseng et al. 2008). The normal cuboidal epithelium had dose-dependently transformed into squamousepithelium, with the most notable change found in the 1,500 mg/kg/day group. Incidence data forhistological lesions were not reported (Tseng et al. 2008). PBDEs 2333. HEALTH EFFECTSSerum T3 levels were significantly reduced by 16% on PND 21 in male offspring of rat dams exposed to decaBDE at 146 mg/kg/day via gavage from GD 6 to PND 21; no change was observed in serum T3 levels in female offspring or serum T4 or TSH levels in either sex at doses ≤146 mg/kg/day (Fujimoto et al. 2011). Following maternal exposure to decaBDE at 0, 10, 500, or 1,500 mg/kg/day via gavage on GDs 0–17, serum T3 levels were significantly decreased by 21% in the 10 and 1,500 mg/kg/day groups in male mouse offspring examined on PND 71 (female offspring not examined); serum T3 levels were notsignificantly altered in the 500 mg/kg/day group and serum T4 levels were not altered in any group (Tsenget al. 2008). In neonatal male mice exposed to 0, 6, or 20 mg/kg/day from PND 2 to 15, serum T4 was reduced by ~8 and 22% at 6 and 20 mg/kg/day, respectively (Rice et al. 2007). This finding was reportedas a dose-related trend; however, pair-wise statistics were not reported. No exposure-related changes inserum T4 levels were observed in similarly exposed neonatal female mice (Rice et al. 2007). Serum T3 levels were not examined by Rice et al. (2007).Immune System DevelopmentLower-brominated PBDEs: In a one-generation study, F0 rats were exposed to pentaBDE at 0, 0.5, 5, or25 mg/kg/day via gavage for 70 days prior to mating, through mating, gestation, and lactation (PND 21)(Bondy et al. 2013). F1 rats continued exposure to their respective doses from PND 22 to 42. Half of theF1 rats were sacrificed on PND 43 and assessed for serum immunoglobin levels, B and T lymphocytequantification in the spleen, spleen cell proliferation in vitro, and immune organ weight and histology. The remaining F1 rats were assessed for immune function at PND 56 using the KLH antigen immunechallenge. Serum IgE and IgG1 levels were significantly reduced by 77 and 53%, respectively, in females at 25 mg/kg/day; no changes were observed in serum IgE or IgG1 levels in males or serum IgM, IgA, IgG2a, IgG2b, or IgGc levels in either sex. All exposed groups showed a significant, dose-relatedreduction in the proportion of B cells and a significant concomitant increase in the proportion of T cells in the spleen (7–18%). In vitro, increased proliferation of unstimulated spleen cells was observed in cellsharvested from F1 males and females from the 25 mg/kg/day group; however, the proliferative response to ConA or lipopolysaccharide (LPS) stimulation was not affected by pentaBDE exposure. In the thymiccortex from F1 rats, mild increases in apoptotic lymphocytes and tingible macrophages were observed inF1 males (0/14, 1/19, 3/13, and 3/13) and F1 females (0/13, 3/17, 0/13, and 4/14) in control, 0.5, 5, and 25 mg/kg/day groups, respectively. Trend test analysis indicated that the increase was significantly dose-related in males, but not females. No treatment-related histopathological changes were observed in the Peyer’s patches, mesenteric lymph nodes, or spleen of F1 rats. Absolute thymus weights were significantly increased in the 5 mg/kg/day males and females and relative thymus weight was only PBDEs 2343. HEALTH EFFECTSincreased in the 5 mg/kg/day females; no changes were observed in thymus weights at 25 mg/kg/day. These organ weight changes reflect observed body weight changes (significantly increased in the5 mg/kg/day group). In the immune challenged rats, no exposure-related changes were observed inKLH-specific IgG levels or changes in delayed-type hypersensitivity responses to KLH injections (Bondyet al. 2013). In another one-generation study (pre-mating day 28 to PND 21), maternal exposure to pentaBDE-dose vanilla wafers at doses up to 11.4 mg/kg/day had no effect on thymus weight in PND 21 offspring (Poon et al. 2011).In a poorly-reported study, mouse dams were exposed to pentaBDE at 0, 50, 100, or 200 mg/kg/day viagavage from GD 6 to PND 21 (Hong et al. 2010). In PND 21 offspring, study authors report exposure-related decreases in serum IgM and IgG1, decreased absolute and relative spleen weights at≥100 mg/kg/day, and reduced cellularity levels at ≥100 mg/kg/day. No statistically significant increasesin T- or B-cell lymphocyte proliferation were observed in vitro. The only significant effect observed in PND 63 offspring was increased T-cell proliferation following in vitro exposure to ConA at200 mg/kg/day.In another study, absolute spleen weight was significantly increased by 9% in PND 140 male offspring ofrat dams exposed to single pentaBDE doses of 0.06 and 0.3 mg/kg on GD 6; relative spleen weight wasonly significantly increased in the 0.06 mg/kg group (12%) (Kuriyama et al. 2005). No organ weightchanges were observed in the thymus of PND 140 male rats (Kuriyama et al. 2005). Following maternalexposure to pentaBDE at 0 or 18 mg/kg/day via gavage from GD 6 to PND 18, no exposure-relatedchanges were observed in spleen or thymus weights measured in rat offspring on PNDs 3, 12, 18, and 31 (Ellis-Hutchings et al. 2006).DecaBDE: Immune function in PND 28 mice was assessed using the RSV intranasal infection testfollowing maternal exposure to decaBDE from GD 10 to PND 21 at doses of 0, 3.3, 34, 260, or3,100 mg/kg/day (Watanabe et al. 2008). Typical features of pneumonia due to RSV infection wereobserved in all RSV-infected mice; however, exacerbation of histopathological changes in the lung wasobserved in 50% of mice exposed to 3,100 mg/kg/day, including hypertrophy and/or hyperplasia. Fivedays after the RSV infection, pulmonary viral titers of RSV and bronchoalveolar lavage fluid (BALF)levels of IFN-γ were significantly increased in the 260 mg/kg/day group (titers and BALF fluid notassessed in the 3,100 mg/kg/day group). Additionally, mRNA expression of RANTES (a characteristic marker of severity of inflammation in the lungs due to an RSV infection) was significantly elevated at34 and 260 mg/kg/day (Watanabe et al. 2008). In a second study, immune function in PND 28 mice was PBDEs 2353. HEALTH EFFECTSalso assessed using the RSV infection test following maternal exposure to decaBDE from GD 10 toPND 21 at doses of 0, 290, or 2,900 mg/kg/day (Watanabe et al. 2010b). In this study, pulmonary viraltiters of RSV were significantly increased in the 2,900 mg/kg/day group 1 and 5 days post-infection. BALF levels of TNF-α, IL-6, and IL1β were significantly decreased one day post-infection (cytokinelevels at 5 days postexposure were not reported). No exposure-related changes were observed in BALFcell TNF-α production in vitro in response to LPS. In a poorly-reported study, mouse dams were exposed to decaBDE at 0, 500, 2,500, or 12,500 mg/kg/dayvia gavage from GD 6 to PND 21 (Hong et al. 2010). Absolute and relative thymus weights weresignificantly increased in offspring the 2,500 mg/kg/day group at PND 21, but not the 500 or12,500 mg/kg/day groups. At PND 21, no exposure-related changes were reported for spleen weight,spleen, or thymus cellularity, serum IgM or IgG1 levels, or T- or B-cell lymphocyte proliferation in vitro. The study authors report an increase in the relative B cell population and a decrease in the relative distribution of macrophage cells in spleens from pups exposed to decaBDE (no statistics provided). Noexposure-related effects were observed at PND 63. In another study, no exposure-related changes inspleen weight were observed in PND 71 male offspring of mouse dams exposed to decaBDE at doses upto 1,500 mg/kg/day from GD 0 to 17 (Tseng et al. 2008).SummaryTeratology, fetotoxicity, and physical growth and development: No human studies have evaluatedassociations between embryotoxicity or fetotoxicity and PBDE exposure. Evidence for altered physicalgrown and development from human studies is inconsistent. Available data from animal studies do notindicate that PBDEs are embryotoxic or fetotoxic at PBDE doses below doses that elicited maternaltoxicity, although occasional observations of reduced pup weight were reported. Taken together, it isunlikely that oral PBDE exposure will cause embryotoxicity or fetoxicity in humans; however, dataindicate that PBDE exposure could potentially lead to low birth weight.Neurodevelopment: Evidence from both human and animal studies indicates that oral PBDE exposurecan lead to adverse effects in neurodevelopment, leading to altered neurobehavior later in life.Reproductive system development: Based on limited human and adequate animal data, it is possible thatthat oral PBDE exposure during development may adversely affect the developing reproductive system, PBDEs 2363. HEALTH EFFECTSparticularly the male reproductive system. However, data are too limited to adequately determine whether or not PBDE exposure in infants and children will lead to altered reproductive performance.Endocrine system development: While human data are inconsistent, they suggest that PBDEs can interactwith thyroid hormone homeostasis in infants and children. These data, along with available animalstudies, indicate that the thyroid is a target of concern for PBDE exposure, especially lower-brominated PBDEs.Immune system development: Animal data suggest that oral PBDE exposure during development maylead to immunosuppression; however, data are too limited to adequately assess the immunotoxic potentialof PBDE exposure in infants and children.The highest NOAEL values and all LOAEL values from each reliable study for developmental effects ineach species and duration category are recorded in Table 3-2 and plotted in Figure 3-2.3.2.2.7 CancerHuman Studies. There was no clear association between risk of non-Hodgkin’s lymphoma (NHL) and exposure to 2,2'4,4'-tetraBDE in a case-control study of 77 Swedish men and women who were recruitedin 1995–1997 and ranged in age from 28 to 85 years (Hardell et al. 1998; Lindstrom et al. 1998). Adiposetissue concentrations of 2,2',4,4'-tetraBDE (BDE 47) (used as a marker for total PBDE exposure) were compared in 19 patients with NHL, 23 patients with malignant melanoma, 8 patients with other cancers orin situ changes, and 27 persons with no cancer diagnosis. The highest concentrations were seen in thepatients with NHL. The mean concentration of BDE 47 was 13.0 ng/g (ppb) lipid (range 1.0–98.2 ppb) in the 19 NHL patients and 5.1 ppb (range 0.6–27.5 ppb) in the 27 persons without known malignancies. Logistic regression, adjusted for age, gender, sum of PCBs, and sum of chlordanes, was performed on cases and controls in three concentration groups (<2.05, 2.05–<5.43, and ≥5.43 ppb). A nonsignificantlyelevated risk with a suggestive dose-response was found for NHL in the two highest concentration groupscompared with the lowest group; the ORs and 95% CIs were 1.9 (0.3–14) and 3.8 (0.7–26) in the middleand high groups, respectively. Although the risk was highest in the group with the highest concentrationof 2,2'4,4'-tetraBDE (p=0.09 for trend), there was no significant difference between cases and controls (p=0.14). The results for patients with malignant melanoma did not differ from controls. PBDEs 2373. HEALTH EFFECTSHardell et al. (2006) found no association between testicular cancer risk and serum PBDE (sum ofcongeners 47, 99, and 153, but concentrations not reported) in a study of 58 cases recruited from hospitalsin Sweden and age-matched controls. Blood was also collected from the mothers of the cases andcontrols (44 of the case mothers and 45 of the control mothers agreed to participate). There was a marginally significant relationship between maternal PBDE and the risk of testicular cancer in sons (OR2.5, 95% CI 1.02–6.0 using median concentration in mothers of controls as a cut-off) and some evidence of a dose-response (OR 3.2, 95% CI 1.1–11 for those with high concentrations of PBDE in blood [>75th percentile] and OR 1.8, 95% CI 0.6–7.9 for those with low concentrations of PBDE in blood [median–75th percentile]). The relevance of this finding is uncertain, as it is unclear the extent to which the case mothers’ body burden of PBDE at the time of the study might relate to body burden when givingbirth to the cases (approximately 30 years previously, as median age of the cases was 30 years). A case-control study of Singaporean males found no significant association between serum levels of BDE 47 and risk of prostate cancer (Pi et al. 2016). The study included 240 prostate cancer incident cases and268 controls. Serum concentrations of BDE 47 were 37 ng/g lipid in cases and 58 ng/g lipid in controls. Other BDEs analyzed in serum included BDE 99, BDE 100, BDE 153, BDE 154, and BDE 183, but theconcentrations were below the limit of detection of the analytical method (0.3–20 pg/g lipid). PBDE concentrations in adipose tissue (sum of congeners 28, 47, 66, 100, 99, 85, 154, 153, 138, and 183)were significantly higher in 21 cases with exocrine pancreatic cancer recruited in Sweden 1996–1999(median 3.1 ng/g lipid) than in 59 controls comprising 20 males undergoing transurethral resection forbenign prostate hyperplasia and 39 females undergoing hysterectomy 1997–1998 in the samegeographical area (median 1.6 ng/g lipid) (Hardell et al. 2007). Case-control analysis found that the riskof pancreatic cancer was not significantly increased with lipid PBDE (OR 3.90, 95% CI 0.93–16.3) usingmedian concentration in controls as a cut-off after adjustment for age, sex, and BMI at tissue sampling,but that the increase in risk was significant when the BMI adjustment was performed for the year before tissue sampling (OR 7.67, 95% CI 1.53–38.5, body weight 1 year before tissue sampling obtained byquestionnaire).Serum concentrations of PBDEs (10 tri- to hepta-BDEs) were not associated with risk of thyroid cancer ina nested case-control study in the Prostate, Lung, Colorectal, and Ovarian Cancer Screening Trial, a large multicenter clinical trial in the United States (Aschebrook-Kilfoy et al. 2015). The study included 104 case of thyroid cancer and 208 controls matched to cases by race, sex, birth date, center, and blood collection date. Median lipid adjusted concentrations of ΣPBDEs (sum of BDE 47, 99, 100, and 153)were 12.8 ng/g for cases and 19.4 ng/g for controls. For ΣPBDEs, the OR for the fourth versus the first PBDEs 2383. HEALTH EFFECTSquartile was 0.62 (95% CI 0.29–1.30). Restricting the analyses to cases with the papillary subtype (75%of the cases) did not alter the findings. A case-control study found no evidence of an association between adipose concentrations of PBDE (BDE47, BDE 99, BDE 100, BDE 153, BDE 154, and their sum) and breast cancer risk in women from the San Francisco Bay area of California (Hurley et al. 2011). The study population included 78 cases with histologically confirmed invasive breast cancer and 56 controls with benign histological changes undergoing surgical breast biopsies. PBDE concentrations in both cases and controls were relatively high(median values of 56.32 and 72.25 ng/g lipid in cases and controls, respectively, for total BDE). Anothercase-control study of 170 native Alaskan women reported that cases (n=75) had significantly higherserum levels of BDE 47 (38.8 ng/g lipid) than controls (n=95, 25.1 ng/g lipid) (Holmes et al. 2014). Theresult of univariate logistic regression analysis showed increase risk that approached statistical significance (OR 1.79, 95% CI 0.97–3.32); no statistical significance was apparent in multivariableanalysis (OR 1.58, 95% CI 0.75–3.33). Animal Studies. Information on carcinogenic effects of PBDEs in animals is limited to results of chronicbioassays of decaBDE mixtures in rats and mice (Kociba et al. 1975; Norris et al. 1975a; NTP 1986). Assummarized below, these studies provide limited evidence for the carcinogenicity of decaBDE in animals. No carcinogenicity studies of octaBDE or pentaBDE were located in the available literature.NTP evaluated the carcinogenicity of commercial-grade decaBDE (94–97% pure, no detected brominated dioxins or furans) in Sprague-Dawley rats (50/sex/dose) and B6C3F1 mice (50/sex/dose) that wereexposed in the diet for 103 weeks and observed for an additional 0–1 weeks (NTP 1986). Comprehensivegross and histological examinations were performed on all animals in all dose groups including those thatwere moribund or died during the study. Reported estimated dose levels in the rats were 1,120 and 2,240 mg/kg/day in males and 1,200 and 2,550 mg/kg/day in females. Incidences of liver neoplastic nodules in low- and high-dose male rats (7/50 and 15/49, respectively) and high-dose female rats (9/50)were significantly greater than in controls (1/50 in both males and females) (p≤0.03, Fisher Exact test)and showed positive dose-related trends (p<0.001, Cochran-Armitage trend test). Incidences ofhepatocellular carcinoma alone (1/50, control males; 1/50, low-dose males; 1/49, high-dose males; 0/50, control females; 2/49, low-dose females; and 0/50, high-dose females) were not significantly increased inthe treated rat groups compared to controls. The increased incidences of neoplastic nodules were considered as “some evidence of carcinogenicity” in both sexes. However, although it was concluded that there was some evidence of carcinogenicity in male and female rats based on “neoplastic nodules,” PBDEs 2393. HEALTH EFFECTSthis is a poorly defined and understood term that is no longer used by NTP to characterizehepatoproliferative lesions in rats. A dose-related trend for mononuclear cell leukemia was observed in treated male rats but was not considered to be biologically significant because of a high incidence incontrol animals.Reported estimated doses in the mice were 3,200 and 6,650 mg/kg/day in males and 3,760 and 7,780 mg/kg/day in females (NTP 1986). Hepatocellular adenoma or carcinoma (combined) occurred atsignificantly increased incidences in low-dose male mice (22/50, p=0.002) and high-dose male mice (18/50, p=0.019) in comparison to controls (8/50) and showed a positive dose-related trend (p=0.021). Incidences of hepatocellular carcinoma alone were not significantly increased in either the low- or high-dose male mice. Slightly elevated incidences of thyroid gland follicular cell adenoma or carcinoma (combined) were additionally observed in exposed male mice but the increases were not statisticallysignificant (control, 0/50; low dose, 4/50; high dose, 3/50). Incidences of follicular cell hyperplasia were significantly increased in male mice as summarized in the subsection on Endocrine Effects inSection 3.2.2.2. No significantly increased incidences of neoplastic lesions were observed in the female mice. NTP (1986) concluded that the significant increase in liver tumors and equivocal increase inthyroid tumors represented equivocal evidence of carcinogenicity in male mice. The evidence ofcarcinogenicity in the male mice was considered limited by an early loss of control animals. Losses ofcontrol male mice were significant during the first year of the study, but were subsequently comparable tothe dosed mice; the early losses were presumed to be due to fighting among animals in both control and treatment groups.The carcinogenicity of decaBDE was also evaluated in Sprague-Dawley rats (25/sex/dose) that were exposed to dietary doses of 0, 0.01, 0.1, or 1.0 mg/kg/day for approximately 2 years (702 days for males,735 days for females) (Kociba et al. 1975; Norris et al. 1975a). The commercial mixture was comprisedof 77.4% decaBDE, 21.8% nonaBDE, and 0.8% octaBDE and therefore differs from typical decaBDEformulations containing ≥97% decaBDE. Comprehensive histological examinations showed no exposure-related neoplastic effects. The ability of this study to detect carcinogenic changes is limited bythe very low dose levels in comparison to those tested in the NTP (1986) bioassay.Summary. With the exception of one small case-control study reporting possible associations betweenadipose PBDE concentrations and risk of pancreatic cancer, there is no evidence for carcinogenicity ofPBDEs in human studies. There is limited evidence of carcinogenicity in animals in a NTP bioassay withdecaBDE (significantly increased incidences of neoplastic liver nodules in rats and combined PBDEs 2403. HEALTH EFFECTShepatocellular adenomas and carcinomas in mice). The Cancer Effect Levels (CELs) for decaBDE in the NTP (1986) study are recorded in Table 3-3 and plotted in Figure 3-3.3.2.3 Dermal Exposure A few studies have examined groups of chemical workers involved in the manufacture and distribution ofPBDEs (Bahn et al. 1980; Brown et al. 1981; Chanda et al. 1982; Landrigan et al. 1979; Rosenman et al. 1979; Stross et al. 1981). Although the route of exposure (inhalation relative to dermal) of these workers has not been well defined, they appear to have had a high potential for dermal exposure (Anderson et al.1978). Results from these studies are discussed in this section, as well as in Section 3.2.1. Dermalexposure may not be an important route of concern for PBDEs because dermal absorption is likely to be low, particularly for the highly brominated congeners, based on in vitro dermal absorption assays (Hugheset al. 2001; Roper et al. 2006).3.2.3.1 Death No reports of death in humans after dermal exposure to PBDEs were located in the available literature.No deaths occurred in rabbits that were observed for 14 days following a single ≤2,000 mg/kg dermaldose of decaBDE, octaBDE, or pentaBDE (IRDC 1974, 1975a, 1975b). The PBDEs were applied to clipped intact skin, covered with an occlusive barrier, and washed from the treatment site 24 hours later. 3.2.3.2 Systemic EffectsNo studies were located regarding respiratory, cardiovascular, gastrointestinal, hematological,musculoskeletal, hepatic, or renal effects in humans or animals after dermal exposure to PBDEs.Systemic effects that have been observed in humans and animals following dermal exposure to PBDEs are described below. The highest NOAEL and all LOAEL values from each reliable study for systemic end points in each species and duration category are recorded in Table 3-4.Endocrine Effects. There is suggestive evidence of hypothyroidism in a small group of workers who were occupationally exposed to decaBDE (Bahn et al. 1980) as summarized above and detailed in Section 3.2.1.2.Table 3-4 Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - Dermal Exposure/ LOAEL Species (Strain) Duration/ Frequency (Route) System NOAEL Reference Chemical Form Comments Less Serious Serious ACUTE EXPOSURE Systemic Rabbit 24 hr (New Zealand) Bd Wt 2000 mg/kg IRDC 1975a OctaBDE (technical) 21032103 mg/kgRabbit 24 hr (New Zealand) Bd Wt 2000 mg/kg IRDC 1975b PentaBDE (technical) 21052105 mg/kg3. HEALTH EFFECTSPBDEs241Note on chemical form: The chemical forms included technical octaBDE and pentaBDE mixtures (exact compositions were not reported). Bd Wt = body weight; hr = hour(s); LOAEL = lowest-observed-adverse-effect level; NOAEL = no-observed-adverse-effect level PBDEs 2423. HEALTH EFFECTSNo studies were located regarding endocrine effects in animals after dermal exposure to PBDEs.Dermal Effects. No studies evaluating dermal exposure of PBDEs were located.There was no evidence of primary irritation in intact skin of rabbits that were dermally exposed to aformer commercial decaBDE mixture (500 mg as dry solid was applied to clipped skin and occluded for24 hours) (IRDC 1974). A similar application of 77.4% decaBDE (containing 21.8% nonaBDE and 0.8%octaBDE) (dry solid, amount not reported), octaBDE (500 mg as dry solid), or pentaBDE (0.5 mL as a viscous liquid) was also non-irritating to intact rabbit skin (IRDC 1975a, 1975b). OctaBDE and pentaBDE were non-sensitizing in maximization tests in guinea pigs (Microbiological Associates Inc. 1996). The induction doses consisted of three pairs of interscapular region intradermalinjections of (1) a 50:50 solution of Freund’s adjuvant and corn oil, (2) 2.5% octaBDE or 5% pentaBDEsolutions in corn oil, and (3) 2.5% octaBDE or 5% pentaBDE in the 50:50 corn oil/Freund’s adjuvantsolution. Control groups received the same regimen without PBDEs. After 7 days, the PBDE-treatedanimals received topical applications of neat octaBDE or pentaBDE on the previously treatedinterscapular sites. Two weeks later, the animals were challenged with topical doses of neat octaBDE orpentaBDE on the left flank. Subsequent examination of the test sites at 24, 48, 72, 96, or 120 hours afterthe challenge dose showed no erythema or edema responses in any of the animals, indicating that the PBDEs did not cause delayed contact hypersensitivity.A 10% chloroform solution of 77.4% decaBDE (containing 21.8% nonaBDE and 0.8% octaBDE) did notinduce bromacne when applied to the ear of rabbits for 30 days (Norris et al. 1975a). A slight erythematous response and slight exfoliation were the only observed effects. No additional information was reported on the design and results of this acnegenesis study.Ocular Effects. No studies were located regarding ocular effects in animals after dermal exposure toPBDEs.Ocular effects were investigated in rats that had 100 mg decaBDE (solid), 100 mg octaBDE (solid), or0.1 mL pentaBDE (viscous liquid) instilled into the conjunctival sac (IRCD 1974, 1975a, 1975b). Theeyes were examined for irritation after 24, 48, and 72 hours and 7 days and corneal injury after 72 hours. There were no exposure-related effects with decaBDE or octaBDE, although pentaBDE caused slightevidence of corneal damage in one of six rats (IRDC 1975b). PBDEs 2433. HEALTH EFFECTSBody Weight Effects. No studies were located regarding body weight effects in humans after dermalexposure to PBDEs.There were no adverse effects on body weight in rabbits that were observed for 14 days following a single≤2,000 mg/kg dermal dose of decaBDE, octaBDE, or pentaBDE (IRDC 1974, 1975a, 1975b). ThePBDEs were applied to clipped intact skin, covered with an occlusive barrier, and washed from the treatment site after 24 hours.No studies were located regarding the following effects in humans or animals after dermal exposure toPBDEs:3.2.3.3 Immunological and Lymphoreticular Effects 3.2.3.4 Neurological Effects 3.2.3.5 Reproductive Effects3.2.3.6 Developmental Effects 3.2.3.7 Cancer3.3 GENOTOXICITYLimited information exists regarding the in vivo genotoxicity of PBDEs. The frequency of micronuclei in peripheral lymphocytes was significantly higher in 23 Chinese workers who dismantled electronic waste(median PBDE serum concentrations 382 ng/g lipid weight) compared with 26 unexposed workers (median PBDE serum concentrations 158 ng/g lipid weight) (Yuan et al. 2008). Multivariate logistic regressionanalysis of several risk factors showed that only history of engaging in dismantling electronic waste was a significant predictor of micronuclei frequencies. The investigators also noted that pollutants other than PBDEs also may have played a role in the increased frequency of micronuclei. No evidence of oxidativeDNA damage was found in peripheral lymphocytes from exposed workers. A study of 84 healthy Korean subjects from the general population reported that serum concentrations of BDE 47 and BDE 99 were notassociated with telomere length in peripheral lymphocytes (Shin et al. 2010); actual concentrations ofPBDEs were not provided. Telomers are complex structures consisting of repeat DNA sequences and associated proteins located at the end of chromosomes which protect chromosomes from end-to-end fusions. Dosing of mice with up to 1,250 mg/kg/day pentaBDE (mixed congeners) by gavage once per day for3 consecutive days did not increase the frequency of micronuclei in blood or bone marrow cells (Witt et al. 2008). In another gavage study, dosing of pregnant mice with ≥10 mg/kg/day decaBDE (the lowest dose PBDEs 2443. HEALTH EFFECTStested) on GDs 0–17 resulted in significant sperm chromatin DNA damage in male offspring examined at71 days of age, as indicated by DNA denaturation induction and increased DNA fragmentation index in the sperm chromatin structure analysis (Tseng et al. 2013). The investigators suggested that hydrogenperoxide, which was increased in sperm cells, may have been involved in induction of oxidative DNAdamage. A summary of in vivo genotoxicity studies of PBDEs is presented in Table 3-5.Results from in vitro assays for gene mutation in various Salmonella typhimurium strains and inEscherichia coli WP2 uvrA conducted with 2,2’,4,4’,5-pentaBDE and decaBDE yielded negative results in the presence or absence of metabolic activation (Evandri et al. 2003; NTP 1986). Assays conducted in mammalian cells yielded mixed results. Tests of decaBDE (BDE 209) for sister chromatid exchange and chromosomal aberrations in Chinese hamster ovary cells were negative and the same was reported forgene mutation in mouse lymphoma L5178Y cells (NTP 1986). These tests were conducted with and without metabolic activation. However, decaBDE induced DNA damage (Comet assay) in humanSK-N-MC neuroblastoma cells (Pellacani et al. 2012) and SW 480 colon carcinoma cells (Curcic et al.2014). Several congeners (BDE 47, 99, 253, 183, and 209) induced micronuclei in human mammarycarcinoma cells (Barber et al. 2006). Both BDE 32 and BDE 47 caused DNA damage (Comet assay) in HepG2 hepatocellular carcinoma cells (Saquib et al. 2016). The congener 2,2',4,4'-tetraBDE (BDE 47) also induced micronuclei formation and DNA damage (Comet assay) in human neuroblastoma cells (Gaoet al. 2009; He et al. 2008a; Pellacani et al. 2012) and DNA damage (Comet assay) in rat primaryhippocampal neurons (He et al. 2008b). However, BDE 47 did not cause DNA damage in normal human hepatocytes, as assessed by the Comet assay (An et al. 2011). BDE 47 also induced gene recombination in Chinese hamster SPD8/V79 cells, but not Sp5/V79 cells (Helleday et al. 1999). In the same study,3,4-diBDE (BDE 12) and 2-monoBDE (BDE 1) produced positive results for gene recombination in both Chinese hamster SPD8/V79 cells and Sp5/V79 cells (Helleday et al. 1999). Experiments conducted by Jiet al. (2011) in chicken DT40 cell lines showed that BDE 47 and 2,2,4’,5-tetraBDE (BDE 49) couldinduce DNA damage (double strand breaks; identified by γ-H2AX focus formation) and that BDE 47 could also induce chromosomal aberration. Ji et al. (2011) also reported that tetraBDEs had a greatergenotoxic potential than PBDEs with a higher number of bromine substitutes and that hydroxylated analogs of tetraBDEs were more genotoxic than tetraBDEs. These investigators suggested that DNAdamage caused by tetraBDEs and hydroxylated analogs is mediated through ROS, which leads to replication blockage and subsequent chromosomal breaks. However, Song et al. (2009) did not find evidence of DNA damage (Comet assay) in human adreno cortical carcinoma cells exposed to the hydroxylated metabolite OH-BDE-47 or OH-BDE-85. DNA damage by PBDEs could be mediated viacovalent binding of PBDE quinone metabolites, as both Lai et al. (2011) and Huang et al. (2015) have PBDEs 2453. HEALTH EFFECTSTable 3-5. Genotoxicity of Polybrominated Diphenyl Ethers (PBDEs) In VivoSpecies (test system) Compound End point Results ReferenceHuman peripheral Total PBDEs Micronucleous assay ± Yuan et al. 2008lymphocytesHuman peripheral Total PBDEs Oxidative DNA damage − Yuan et al. 2008lymphocytesHuman peripheral BDE 47 Telomere length − Shin et al. 2010lymphocytesHuman peripheral BDE 99 Telomere length − Shin et al. 2010lymphocytesMouse blood and bone PentaBDE Micronucleous assay − Witt et al. 2008marrow (unspecified)Mouse sperm BDE 209 DNA damage (sperm + Tseng et al. 2013chromatin structure analysis)+ = positive result; – = negative result; ± = inconclusive result; BDE = brominated diphenyl ether PBDEs 2463. HEALTH EFFECTSreported that PBDE quinone metabolites covalently bind to DNA in vitro to form DNA adducts. Thesemetabolites would result from oxidation of hydroxylated metabolites to dihydroxylated metabolites bymicrosomal cytochrome P450s. A summary of the in vitro genotoxicity data for PBDEs is presented inTable 3-6.Overall, the information available from in vivo and in vitro studies with a few PBDE congeners is insufficient to make generalizations regarding the genotoxicity of PBDEs.3.4 TOXICOKINETICS3.4.1 Absorption 3.4.1.1 Inhalation Exposure No studies were located regarding absorption of PBDEs in humans after inhalation exposure. Evidence for the inhalation absorption of lower-brominated PBDEs in animals was provided byobservations of systemic toxicity in rats that were intermittently exposed to a commercial octaBDEproduct (bromine content 78.7%) as dust aerosol for 13 weeks (Great Lakes Chemical Corporation 2000). The absorption of the lower-brominated BDE congeners was indicated by the occurrence of hepatic,thyroid, and ovarian effects in rats following exposure to 16 or 202 mg/m3 for 6 hours/day, 5 days/week,for 13 weeks.No studies were located that quantified absorption of inhaled PBDEs, but Staskal et al. (2005) reported absorption efficiencies of 91% in mice given single intratracheal doses of 14C-BDE 47 in corn oil. Thisestimate was derived by comparing the radioactivity profiles in urine, feces, and tissues 5 days afterintratracheal or intravenous administration of 1 mg/kg doses of 14C-BDE 47.3.4.1.2 Oral ExposureHuman Data. No information was located regarding absorption of PBDEs in humans followingcontrolled oral exposure. PBDEs 2473. HEALTH EFFECTSTable 3-6. Genotoxicity of Polybrominated Diphenyl Ethers (PBDEs) In VitroResultsWith WithoutSpecies (test system) Compound End point activation activation ReferenceProkaryotic organisms:Salmonella BDE 209 Gene mutation – – NTP 1986typhimurium, TA98, TA100, TA1535,TA1537S. typhimurium, TA98,TA100BDE 99 Gene mutation – – Evandri et al. 2003Escherichia coli, WP2uvrABDE 99 Gene mutation – – Evandri et al. 2003Mammalian cells:Human L02 normalhepatocytesBDE 47 DNA damage (Comet assay)No data – An et al. 2011Human MCF-7 mammary carcinoma cellsBDE 47, 99,153, 183, 209Micronuclei No data + Barber et al. 2006Human SW 480 colon carcinoma cellsBDE 209 DNA damage (Comet assay)No data + Curcic et al. 2014Human SH-SY5Y neuroblastoma cellsBDE 47 DNA damage (Comet assay)No data + Gao et al. 2009Human SH-SY5Y BDE 47 Micronuclei No data + He et al. 2008aneuroblastoma cellsHuman SH-SY5Y neuroblastoma cellsBDE 47 DNA damage(Comet assay)No data + He et al. 2008aHuman SK-N-MCneuroblastoma cellsBDE 47 DNA damage(Comet assay)No data + Pellacani et al. 2012Human SK-N-MCneuroblastoma cellsBDE 209 DNA damage(Comet assay)No data + Pellacani et al. 2012Human HepG2 hepatocellularcarcinoma cellsBDE 32, 47 DNA damage (Comet assay)No data + Saquib et al. 2016 Human H295Radrenocorticalcarcinoma cellsOH-BDE-47OH-BDE-85DNA damage(Comet assay)NA – Song et al. 2009Rat primaryhippocampal neuronsBDE 47 DNA damage(Comet assay)No data + He et al. 2008bMouse lymphoma L5178Y cellsBDE 209 Gene mutation – – NTP 1986Chinese hamsterSp5/V79 cellsBDE 47 Gene recombinationNo data – Helleday et al. 1999Chinese hamsterSPD8/V79 cellsBDE 47 Gene recombinationNo data + Helleday et al. 1999Chinese hamsterSp5/V79 cellsBDE 12 Gene recombinationNo data + Helleday et al. 1999 PBDEs 2483. HEALTH EFFECTSTable 3-6. Genotoxicity of Polybrominated Diphenyl Ethers (PBDEs) In VitroResultsWith WithoutSpecies (test system) Compound End point activation activation ReferenceChinese hamsterSPD8/V79 cellsBDE 12 Gene recombinationNo data + Helleday et al. 1999Chinese hamsterSp5/V79 cellsBDE 1 Gene recombinationNo data + Helleday et al. 1999Chinese hamsterSPD8/V79 cellsBDE 1 Gene recombinationNo data + Helleday et al. 1999Chinese hamster ovarycellsBDE 209 Sister chromatid exchange− − NTP 1986Chinese hamster ovarycellsBDE 209 Chromosomalaberrations− − NTP 1986Other cellsChicken DT40 cell line BDE 47 DNA damage(γ-H2AX focusformation)No data + Ji et al. 2011BDE 49 DNA damage(γ-H2AX focusformation)No data + Ji et al. 2011BDE 47 ChromosomalaberrationsNo data + Ji et al. 2011+ = positive result; C = negative result; BDE = brominated diphenyl ether; DNA = deoxyribonucleic acid; NA = notapplicable PBDEs 2493. HEALTH EFFECTSAnimal DataOverview: Information regarding oral absorption in animals is available from studies of commercialPBDE mixtures and individual 14C-labeled tetra-, penta-, hexa-, and decaBDE congeners. As summarizedbelow, the most recent and best available estimates of oral absorption efficiencies following gavage administration in lipophilic vehicles indicate a range of 70–85% for tetra- (BDE 47), penta- (BDE 99, BDE 100), and hexa- (BDE 153, BDE 154) congeners, and 10–26% for decaBDE (BDE 209).Information on oral absorption of the commercial pentaBDE mixture DE-71 and the commercialoctaBDE mixture DE-79 is available from studies in which male Sprague-Dawley rats were fed diets containing 0 or approximately 32–33 ng/day (≈120 ng/kg/day) of either mixture in peanut oil for 21 days(Hakk et al. 2001; Huwe et al. 2002b, 2007). The doses were designed to mimic environmental exposureconcentrations. Liver, carcass, and feces were analyzed for major congeners in the penta- and octaBDEformulations 24 hours after the final feeding; urine was not evaluated. The study of the pentaBDEmixture (DE-71) assessed the following six congeners: BDE 47, BDE 85, BDE 99, BDE 100, BDE 153, and BDE 154 (Hakk et al. 2001; Huwe et al. 2007). Based on liver, carcass, and unrecovered concentrations of congeners, and assuming that excretion in the urine was negligible, absorption isestimated to have been 44.3% for penta congener BDE 85 and 84.3–92.4% for the other tetra- to hexaBDE congeners. The study of the octaBDE mixture (DE-79) assessed the following eight congeners:BDE 153, BDE 154, BDE 183, BDE 190, an unknown heptaBDE, and three unknown octaBDEs (Huweet al. 2002b). Based on liver, carcass, and unrecovered concentrations of congeners, and assuming thatexcretion in the urine was negligible, absorption is estimated to have been 84.2–95.1% for the hexaBDEs, 68.5–79.1% for the heptaBDEs, and 55.7–83.3% for the octaBDEs. Early studies with 14C-decaBDE (BDE 209) indicated that gastrointestinal absorption efficiency was low(~9 or <1%) in rats (El Dareer et al. 1987; Norris et al. 1973, 1975b; NTP 1986). Following treatmentwith a single 1 mg/kg dose of 14C-decaBDE in corn oil, administered as a low purity commercial mixture (77.4% decaBDE, 21.8% nonaBDE, 0.8% octaBDE) by gavage, 90.6 and >99% of the dose waseliminated in the feces within 24 and 72 hours post-dosing, respectively (Norris et al. 1973, 1975a). An oral absorption efficiency estimate of about 9% is indicated, assuming that fecal radioactivity excreted in24 hours was nonabsorbed, and fecal radioactivity excreted between 24 and 72 hours was from biliaryexcretion of absorbed material. Two feeding studies were conducted in which rats were exposed to acommercial mixture as unlabeled decaBDE (92% pure) on days 1–7 and 14C-decaBDE (98.9% pure) on day 8, followed by unlabelled decaBDE on days 9, 9–10, or 9–11 (El Dareer et al. 1987; NTP 1986). In PBDEs 2503. HEALTH EFFECTSthe first study, dietary concentrations ranged from 238 to 51,100 ppm (six levels) (≈20–4,500 mg/kg/day). Recovery of radioactivity in the feces ranged from 91.3±4.0 to 101±4% of the administered dose and wasnot related to dose level. In the second study, rats were exposed to dietary concentrations of 277 or48,000 ppm (≈20 or 4,300 mg/kg/day). Recovery of radioactivity in the feces ranged from 82.5±4.7 to 86.4±8.5% of the dose and was not related to dose level or time of sacrifice (24, 48, or 72 hours after 14C-decaBDE intake). For both dose levels, the percent of 14C dose remaining in the gut contents (<4%)and gut tissue (<0.04%) decreased with time. Of the radioactivity recovered, >99% was in the feces and gut contents. Based on a comparison of average tissue concentrations following intravenous and oraladministration, NTP (1986) estimated that oral absorption was 0.33±0.19% at the highest dietary level(50,000 ppm).More recent studies indicated that absorption efficiency of 14C-decaBDE (BDE 209) in rats can be influenced by vehicle and have reported absorption efficiencies in the 10–26% range using lipophilicvehicles (Hakk et al. 2002b; Morck and Klasson Wehler 2001; Morck et al. 2003; Riu et al. 2008;Sandholm et al. 2003). In normal and bile duct-cannulated male Sprague-Dawley rats administered single 3-µmol/kg (≈3 mg/kg) doses of 14C-decaBDE (>98% pure) in Lutrol F127/soya phospholipone (34:16, w/w)/water, radioactivity in feces collected for 72 hours accounted for about 90% of the administered dose in conventional rats (Morck and Klasson Wehler 2001; Morck et al. 2003). In bile duct-cannulatedrats, averages of 88 and 9.5% of the dose were recovered in feces and bile, respectively, within 3 days. Radioactivity recovered in urine was <0.1% of the dose in normal and bile-duct cannulated rats. Theradioactivity in bile indicates that at least 10% of the dose was absorbed (Morck and Klasson Wehler2001; Morck et al. 2003). Hakk et al. (2002b) reported similar results in another study in which four bileduct-cannulated male Sprague-Dawley rats were orally administered single 3-µmol/kg doses of14C-decaBDE (>98% pure) in Lutrol 127, soyaphospholipone, and water. Radioactivity in bile and urine collected for 72 hours accounted for 9.2 and <0.1% of the administered dose, respectively. Sandholm etal. (2003) reported an oral bioavailability of 26% for decaBDE from plasma concentration-time curves for144 hours following gavage and intravenous administration of single 2-µmol/kg doses of unlabeleddecaBDE (purity >98%) in a DMA/polyethylene glycol/water vehicle (4:4:1) to Sprague-Dawley rats.DecaBDE concentrations in plasma samples were quantified by gas chromatography/mass spectrometry(GC/MS). Oral bioavailability was calculated by dividing the area under the plasma concentration-timecurve for oral exposure (AUCoral) by the AUCi.v. Qualitative analysis by GC/MS of pooled plasma samples determined 3 major metabolites among 13 hydroxylated metabolites: a hydoxy-octaDDE, ahydroxyl-nonaBDE and a hydoxy-methoxy-hexaBDE. The presence of these phenolic metabolites in the plasma samples indicates that the oral bioavailability (and hence oral absorption efficiency) of decaBDE PBDEs 2513. HEALTH EFFECTSmay have been higher than the calculated value of 26% based on parent-compound plasma concentration-time curves (Sandholm et al. 2003). Riu et al. (2008) reported that >19% of administered radioactivitywas recovered in tissues (including fetuses) and carcasses, 24 hours after oral administration of 2-mg/kg doses of 14C-labeled decaBDE (>99.8% radiopurity) dissolved in peanut oil to pregnant Wistar rats on GDs 16, 17, 18, and 19. This finding indicates an absorption efficiency of about 20%, assuming thatradioactivity recovered in the feces (about 66% of the administered dose) was not absorbed. Because biliary excretion of decaBDE has been demonstrated in rats, this value may underestimate the actual oralabsorption that occurred in this study. However, the aqueous fraction (which would contain hydroxylatedmetabolites and conjugates) accounted for only 4% of the radioactivity in the collected feces, and 97% ofthe radioactivity in the organic-solvent fraction of the feces was determined by high-performance liquid chromatography (HPLC) to be unchanged decaBDE.Estimates of oral absorption efficiencies for BDE 47 in rats and mice have ranged from about 75 to 95%, depending on species and employed measurement techniques (Örn and Klasson-Wehler 1998; Sanders etal. 2006a; Staskal et al. 2005). A single 14.5-mg/kg (30-µmol/kg) gavage dose of 14C-BDE 47 in corn oilwas well absorbed by male Sprague-Dawley rats and male C57Bl mice (Örn and Klasson-Wehler 1998). Approximately 5% of the dose in rats and 7% of the dose in mice was excreted as parent congener in the feces in 24 hours. The investigators concluded that these values represented the non-absorbed doses, indicating that absorption of BDE 47 was 93–95%. Later studies with male F344 rats and male B6C3F1 mice reported oral absorption efficiency estimates for BDE 47 of 75% in rats and 85% in mice (Sanders etal. 2006a). Sanders et al. (2006a) used a more refined technique that compared profiles of radioactivity in urine, feces, and tissues 24 hours after gavage and intravenous administration of single 1-µmol/kg dosesof 14C-BDE 47 in corn oil. Using a similar technique comparing radioactivity profiles in excreta andtissues 5 days after gavage and intravenous administration of single 1-mg/kg doses of 14C-BDE 47 in corn oil, Staskal et al. (2005) reported that oral absorption efficiency was approximately 82% in femaleC57BL/6J mice. Evidence for extensive oral absorption of pentaBDE congeners (BDE 99 and BDE 100) in rats and mice comes from studies that administered single oral doses of about ~28 µmol/kg 14C-BDE 100 (~98% pure) in peanut oil to male Sprague-Dawley rats (Hakk et al. 2006), about 2.2 mg/rat (~15 µmol/kg) of14C-BDE 99 (~98% pure) in corn oil to male Sprague-Dawley rats (Hakk et al. 2002a), or 1 µmol/kg bodyweight of 14C-BDE 99 (~96% pure) in corn oil to male F344 rats and male B6C3F1 mice (Chen et al. 2006). Seventy-two hours after administration of BDE 100, about 73 and 41% of the administered radioactivity remained in tissues of conventional and bile-duct cannulated rats, respectively; fecal PBDEs 2523. HEALTH EFFECTSradioactivity accounted for about 20 and 26% of administered radioactivity in conventional and bile-ductcannulated rats, respectively (Hakk et al. 2006). Assuming that radioactivity excreted in feces within24 hours was non-absorbed, reported values of 11.5% in conventional rats and 16.8% in bile ductcannulated rats indicate oral absorption for BDE 100 of about 89 and 84%, respectively. In the studieswith BDE 99, radioactivity in 24-hour feces accounted for 22.3 and 52.5% of administered radioactivityin conventional and bile-cannulated rats, respectively, indicating absorption efficiencies for BDE 99 of atleast 78 and 48% (Hakk et al. 2002a). Chen et al. (2006) compared radioactivity profiles in 24-hourexcreta and tissues after gavage and intravenous administration of 14C-BDE 99 to arrive at estimated oral absorption efficiencies of approximately 85% in both male rats and male mice. Results from rat and mouse studies with 14C-labeled hexaBDEs (BDE 153 and BDE 154) administered as single doses in oil vehicles also indicate extensive oral absorption (Hakk et al. 2009; Sanders et al.2006b). Hakk et al. (2009) estimated an oral absorption efficiency for BDE 154 of about 77% in maleSprague-Dawley rats, based on radioactivity profiles in 72-hour excreta and tissues after gavage administration of single 11.3-µmol/kg 14C-BDE 154 (>98% pure) to conventional and bile ductcannulated rats. Sanders et al. (2006b) estimated oral absorption efficiencies for BDE 153 of about 70%in male F344 rats and male B6C3F1 mice, based on comparison of radioactivity profiles in 24-hourexcreta and tissues after gavage and intravenous administration of single 1-µmol/kg doses 14C-BDE 153(96% pure). 3.4.1.3 Dermal Exposure No information was located regarding dermal absorption of PBDEs in humans. A dermal absorption efficiency of 62% was reported for female mice exposed to an occluded dermal dose of 1 mg/kg 14C-BDE 47) (Staskal et al. 2005). Profiles of radioactivity in urine, feces, and tissues 5 daysafter dermal or intravenous administration were compared to estimate dermal absorption efficiency.Approximately 15% of the administered dose remained at the site of application 5 days after application. No other in vivo studies of dermal absorption with PBDE congeners or mixtures were located.In in vitro studies with human breast skin and mouse dorsal skin samples exposed to 14C-BDE 147 (radiochemical purity=96.5%) for 24 hours in a flow-through diffusion cell apparatus, reported mean absorption efficiencies (percent of applied radioactivity collected in receptor compartment) were 1.88%for human skin (n=10) and 14.58% for rat skin (n=12) (Roper et al. 2006). PBDEs 2533. HEALTH EFFECTSLimited dermal absorption of 14C-decaBDE by mouse skin was indicated in an in vitro study in which14C-decaBDE dissolved in tetrahydrofuran was applied (three dose levels) to dorsal skin excised fromadult hairless female mice, and fractions of receptor fluid were collected over a 24-hour period (Hughes etal. 2001). Transfer of radioactivity to the receptor fluid was minimal, accounting for only 0.07–0.34% ofthe applied radioactivity. Two to 20% of the radioactivity was found in the skin. The highest percentageof the dose in the skin was associated with application of the lowest dose. Washing the skin with solvent24 hours after application removed 77–92% of the applied dose.3.4.2 Distribution 3.4.2.1 Inhalation Exposure No information was located regarding distribution of PBDEs in humans following controlled inhalation exposure. The distribution of bromine was examined in tissues of rats after inhalation exposure to octaBDE (GreatLakes Chemical Corporation 1978). Groups of rats were exposed to 0, 1.2, 12, 120, or 1,200 mg/m3 ofdusts of octaBDE 8 hours/day for 14 days. At necropsy, sections of the lungs, adipose tissue, and liverwere collected for bromine analysis using a neutron activation technique. The results showedconcentrations of bromine in the lungs and adipose tissue significantly higher in all groups relative to controls; the amounts of bromine detected were concentration-related. In the liver, the concentration ofbromine was also elevated in all groups relative to controls except in the 1.2 mg/m3 exposure group; theelevated concentrations in the liver were not as marked as in the lungs or in adipose tissue.3.4.2.2 Oral ExposureHuman Studies. Overview: No studies were located that examined tissue distribution of PBDEs in humans following controlled oral exposure. Evidence for the transfer of PBDEs from pregnant mothers to the developingfetus and for the transfer of PBDEs from maternal blood to breast milk and then to nursing infants comes from a number of studies of PBDE concentrations in maternal and cord serum samples and breast milksamples from groups of non-occupationally exposed women. Although the contributions of differentpossible exposure routes experienced by these women are uncertain, it is thought that ingestion of PBDEs PBDEs 2543. HEALTH EFFECTSin dust and food represented a significant exposure route. In general, the tetra- and penta-brominated PBDEs have been the predominant congeners detected in maternal and cord serum samples and breastmilk samples, but some recent studies assaying for a wider range of PBDE congeners have found evidence for distribution of hepta-, octa-, or decaBDEs into cord serum and breast milk. In evaluation ofthese studies, the location where the study was performed (e.g., Asia, Europe or North America) is veryimportant, as the patterns of exposure to various congeners were different due to different usage patternsof commercial mixtures (see Section 6.5, General Population and Occupational Exposure, for moreinformation).A number of studies have examined concentrations of PBDEs or hydroxylated PBDEs in maternal and cord blood samples in non-occupationally exposed groups of women from the United States (Chen et al. 2013; Mazdai et al. 2003; Qiu et al. 2009), France (Antignac et al. 2009, 2008), China (Li et al. 2013a), Japan (Kawashiro et al. 2008), Spain (Vizcaino et al. 2011), the Netherlands (Meijer et al. 2008), and Korea (Wan et al. 2010). In the most recent of the U.S. studies, BDE 28, BDE 47, BDE 99, BDE 100, and BDE 153 were detected in 90, 90, 95, 85, and 100% of 20 maternal and 65, 65, 80, 90, and 85% of20 matched cord serum samples, respectively (Chen et al. 2013). Percentages of samples above the limitof detection were lower for BDE 209 (55% maternal and 40% cord) and BDE 154 (25% maternal and 5%cord). Chemical analyses for other PBDE congeners were not conducted in this study. Median concentrations of congeners in maternal serum samples (ng/g lipid) were in the following order: BDE 47 > BDE 153 > BDE 209 > BDE 99 > BDE 100 > BDE 28 > BDE 154. In cord serum samples, the orderwas: BDE 47 > BDE 99 > BDE 209 > BDE 28 > BDE 153 > BDE 100 > BDE 154. Concentrations offour hydoxylated PBDEs (6-OH-BDE-47, 5-OH-BDE-47, 4'-OH-BDE-49 and 6'-OH-BDE-99) were alsodetermined. Geometric mean concentrations (ng/g lipid) of total BDEs and total OH-BDEs in cord samples were higher than those in maternal serum samples (~52% and 42 higher for OH-BDEs and BDEs, respectively). Equal or higher concentrations of total OH-BDEs or total BDEs in cord serum, compared with maternal serum, were found in 85 and 80% of the matched mother-neonate pairs. The results suggest that each of the seven BDEs and four OH-BDEs can cross the placenta and distribute tothe fetus, and that internal exposure via blood is equal to or higher in fetuses, compared with theirmothers. In contrast, BDE 209 was the most abundant congener detected in both maternal and cord bloodsamples from China, where industrial production of BDE 209 may result in exposure (Li et al. 2013a).Similar evidence for transplacental transfer of a set of PBDE congeners more enriched in higherbrominated PBDEs was reported in a recent study that measured concentrations of 19 PBDE congeners in maternal and cord serum samples from 29 mother-neonate pairs from a Wenzhou region of China that is PBDEs 2553. HEALTH EFFECTSthe major region for electronics dismantling and recycling in China (Li et al. 2013a). In both maternaland cord serum samples, the congeners with the highest geometric mean concentrations (ranging from3.32 to 1.78 ng/g lipid in maternal samples) were BDE 207 > BDE 208 > BDE 209 > BDE 28. Geometric means for total cord serum concentrations of lower-brominated congeners (BDE 17 toBDE 154), higher-brominated congeners (BDE 183 to BDE 209), and all BDEs were higher than respective concentrations in maternal blood by about 15, 47, and 41%, respectively (Li et al. 2013a).Frederiksen et al. (2010) utilized a human ex vivo placenta perfusion system to study the kinetics ofplacental transfer of BDE 47, BDE 99, and BDE 209 to the fetus. Placentas were perfused for 4 hours with 1 ng/mL concentrations of the non-labeled congeners. Samples of maternal and fetal compartments were taken during the perfusion, and concentrations in samples were determined by GC/MS. Placentaltransfer of BDE 47 and BDE 99 was demonstrated, and transfer of BDE 47 was faster and more extensive than BDE 99. Transfer of BDE 209 across the placenta to the fetal compartment was not demonstratedwith the detection limits of the techniques employed. Frederiksen et al. (2010) proposed that futureperfusion studies with BDE 209 should use 14C-labeled BDE 209 to increase sensitivity. Evidence for the transfer of PBDEs from maternal blood to breast milk and hence to nursing infantscomes from a number of studies examining PBDE concentrations in breast milk samples. Most studies ofPBDEs in breast milk samples through 2002 measured concentrations of only lower-brominatedcongeners (tetra, penta and hexaBDEs) used in many commercial products up to that time: BDE 47, BDE 99, BDE 100, BDE 153, and BDE 154 (Hites 2004). Additionally, higher brominated PBDEs(e.g., decaBDE; BDE 209) were often omitted from early human biomonitoring due to the inability to accurately measure them at that time. In these early studies, BDE 47 was the congener detected at the highest concentration. A number of more recent studies have included chemical analysis for a greater number of PBDE congeners (including hepta-, octa- and deca-brominated congeners) in breast milksamples collected in the United States (Daniels et al. 2010; Park et al. 2011; Schecter et al. 2010, 2006), France (Antignac et al. 2009, 2008), and Philippines (Malarvannan et al. 2013). The detection of higherbrominated congeners in some of these recent studies indicates that both lower and higher-brominatedcongeners can be distributed to breast milk (Antignac et al. 2008, 2009; Malarvannan et al. 2013; Park etal. 2011; Schecter et al. 2010). Further support for the transfer of PBDEs from mothers to nursing children comes from a report thataverage concentrations of BDE 47 and BDE 99 were statistically significantly (p < 0.05) increased by~5-fold in serum samples from 4-year-old Spanish children (n=202) who had been breastfed, compared PBDEs 2563. HEALTH EFFECTSwith formula-fed 4-year-old children (n=42) (Carrizzo et al. 2007). This study analyzed serum samples for 13 PBDE congeners including tri- (BDE 17, BDE 28), tetra- (BDE 47, BDE 66, BDE 71), penta(BDE 85, BDE 99, BDE 100), hexa- (BDE 138, BDE 153, BDE 154), and hepta-brominated congeners(BDE 183, BDE 190). BDE 47 and BDE 99 were the predominant congeners detected in both breastfed and formula-fed children; concentrations of BDE 47 and BDE 99 in breastfed children were about 16 and 6 times higher than the congener with the next highest concentration, BDE 100. The following congenerswere not detected in the collected serum samples: BDE 17, BDE 28, BDE 66, BDE 71, BDE 85, BDE183, and BDE 190. This study did not control for potential differences in other PBDE exposurepathways, such as ingestion of contaminated dusts or food. Animal Studies. Overview: Tissue distribution studies in animals orally exposed to 14C-labeled BDE 47, BDE 99, BDE 100, BDE 153, BDE 154, and BDE 209 indicate that decaBDE is distributed among tissues somewhat differently than tetra-, penta- and hexaBDEs. While lower-brominated BDE congeners, following absorption and an initial wide distribution, are preferentially accumulated in adipose tissues, absorbed decaBDE is less readily distributed to adipose tissues and appears to preferentially distribute tohighly perfused tissues. Although less likely to partition to adipose tissues, decaBDE was still found in low quantities in adipose tissues in these studies, and has been shown to transfer from dams to fetuses and neonates from exposure during gestational and nursing periods. Two studies with female Sprague-Dawley rats given oral doses of nonlabeled BDE 209 in corn oil from GD 7 to postpartum day (PPD) 4 or8 demonstrated maternal transfer to developing fetuses and neonates by examining whole-body BDE 209concentrations in offspring, but another study with similarly exposed Sprague-Dawley rats found no clearevidence for maternal transfer by examining BDE 209 concentrations in blood from dams and offspring.Results from studies of rats or mice orally exposed to 14C-labeled decaBDE (BDE 209) indicate thatdecaBDE is not as readily distributed to adipose tissues as the lower-brominated BDEs and more readilydistributed to highly perfused tissues (El Dareer et al. 1987; Morck and Klasson-Wehler 2001; Morck etal. 2003; Norris et al. 1975a; NTP 1986; Riu et al. 2008). In male rats administered single gavage dosesof 1 mg/kg of a 14C-labeled commercial decaBDE mixture (77.4% pure containing 21.8% nonaBDE and 0.8% octaBDE), radioactivity could be detected on day 1 in all sampled tissues (adipose, skin, liver, heart, adrenals, spleen, pancreas) (Norris et al. 1975a). On day 16 after dosing, radioactivity was only detected in adrenals and spleen (0.01 and 0.06% of the administered dose per gram of tissue, respectively). In F344 rats fed diets containing a commercial mixture as unlabeled decaBDE (92% pure) on days 1–7, PBDEs 2573. HEALTH EFFECTS14C-decaBDE (98.9% pure) on day 8, and unlabelled decaBDE on days 9, or 9–10, or 9–11, the levels ofradioactivity remaining in tissues 72 hours after the exposure period ended had the following order:gastrointestinal tract > liver > kidney > lung > skin > adipose (El Dareer et al. 1987; NTP 1986). In Sprague-Dawley rats given single 3-µmol/kg (≈3 mg/kg) gavage doses of 14C-decaBDE (>98% pure) in Lutrol F127/soya phospholipone (34:16, w/w)/water, concentrations of radioactivity remaining in tissues3 days after dose administration had the following order (14C nmol equivalents/g lipid concentrations inparentheses): plasma (22 nmol/g) > liver (14.9 nmol/g) > heart ~ small intestine wall ~ adrenal (rangingfrom 5.6 to 4.0 nmol/g) > lung ~thymus ~ kidney (ranging from 2.1 to 1.8 nmol/g) > adipose(0.17 nmol/g) (Morck and Klasson-Wehler 2001; Morck et al. 2003). In pregnant Wistar rats given single2-mg/kg doses of 14C-decaBDE (>99.8% pure) in peanut oil daily on GDs 16–19, concentrations ofradioactivity on a wet weight basis (µg 14C-decaBDE equivalents/g tissue) were highest in adrenals (33 µg/g), ovaries (16 µg/g), and liver (11 µg/g); intermediate in kidneys > stomach > heart ~ placentas >lung > spleen > plasma > uterus > carcass (ranging from 3.90 to 1.11 µg/g); and lowest in adipose tissue (0.79 µg/g), fetuses (0.46 µg/g), brain (0.11 µg/g), and amniotic fluid (0.11 µg/g) (Riu et al. 2008). Results from studies of rats and mice exposed to oral doses of 14C-labeled BDE 47 indicate widedistribution to tissues following absorption with preferential accumulation in fatty tissues, contrasting the preferential distribution of decaBDE to highly perfused tissues (Örn and Klasson-Wehler 1998; Sanders et al. 2006a; Staskal et al. 2005, 2006a). Five days after gavage administration of single 30-µmol/kg doses of 14C-BDE 47 in corn oil to male Sprague-Dawley rats, concentrations of radioactivity had the following order (14C-BDE 47 equivalent nmol/g wet tissue): adipose (706) > lung (12) > kidney (11) >liver (5) > brain (1.9) > plasma (0.76) (Örn and Klasson-Wehler 1998). A similar order of wet tissueconcentrations (nmol/g wet tissue) was observed in male C57Bl mice exposed to 14C-BDE 47: adipose(79) > liver (7) > lung (5) > kidney (3) > brain (1) > plasma (not detected) (Örn and Klasson-Wehler1998). In another study that gave single 1-µmol/kg doses of 14C-BDE 47 to male and female F344 rats incorn oil, the following order of amounts of radioactivity remaining in tissues was observed 24 hours afterdose administration (percentages of administered dose for male rats): adipose (24.6%) > skin (13.0%) >muscle (3.0%) > liver (1.3%) > blood (0.2%) > brain ~ kidney ~ lung (0.1% each) (Sanders et al. 2006a). A similar order was observed in female rats, but the amount of radioactivity in adipose in females (37%)was higher than in male rats (24.6%) (Sanders et al. 2006a). Tissue distribution in male and femaleB6C3F1 mice exposed to 14C-BDE 47 was similar to distribution in F344 rats, with females showinghigher accumulation of radioactivity in adipose than male mice (Sanders et al. 2006a). Staskal et al.(2005) reported similar patterns for tissue distribution in female C57BL/6J mice 5 days afteradministration of single oral doses of 0.1, 1.0, 10, or 100 mg/kg 14C-BDE 47 in corn oil. For all doses, PBDEs 2583. HEALTH EFFECTSthe reported order of remaining radioactivity in tissues was: adipose (8–14% of administered dose) > skinand muscle (~2–3%) > liver (~1%) > brain ~ kidneys ~ lungs (<0.05% each) (Staskal et al. 2005). Thesame tissue distribution pattern was also observed in female C57BL/6J mice given nine 1-mg/kg/day doses of unlabeled BDE 47 in corn oil followed by single 1-mg/kg doses of 14C-BDE 47 (Staskal et al.2006a). Using percent of dose/g of tissue as the dose metric to compare single-dose levels in rats reported by Staskal et al. (2005) with repeated-dose levels in tissues, the tissue concentrations from the repeated-dose scenario with 1 mg/kg were comparable to tissue concentrations from single 10-mg/kg doses and about 2 times greater than concentrations from single 1-mg/kg doses (Staskal et al. 2006a). The results indicate the potential for incremental accumulation of BDE 47 in fat with continued exposure.Results from studies of 14C-BDE 99 (Chen et al. 2006; Hakk et al. 2002a) and 14C-BDE 100 (Hakk et al.2006) indicate an initial wide distribution of penta-brominated congeners to tissues following absorption with preferential accumulation in fatty tissues, similar to results from studies with BDE 47. In male F344 rats given single doses of 0.1 µmol/kg 14C-BDE 99 (~96% pure) in corn oil, levels of radioactivityremaining in tissues 24 hours after dose administration had the following order (percent of administered dose): adipose (20.8%) > skin (7.0%) > muscle (5.2%) > liver (2.1%) > blood = brain (0.3% each) > lung= brain (0.1% each) (Chen et al. 2006). Similar orders of tissue concentrations were seen 24 hours afteroral administration in male rats given single 1, 10, 100, or 1,000 µmol/kg doses, and in female F344 ratsand male and female B6C3F1 mice given single 1 µmol/kg doses (Chen et al. 2006). Preferentialaccumulation in adipose also was found in male Sprague-Dawley rats 24 hours after administration of14.5 µmol/kg doses of 14C-BDE 99 (>98% pure) in corn oil (Hakk et al. 2002a). In male Sprague-Dawleyrats 72 hours after administration of single 25.2 µmol/kg doses of 14C-BDE 100 (>95% pure) in peanutoil, the adipose, gastrointestinal tract, and skin had the highest concentrations of radioactivity(>35 nmol/g fresh tissue weight), liver and lung had intermediate concentrations (~10–15 nmol/g), and other tissues, including lung, adrenals, testes, and muscle, had the lowest concentrations (<10 nmol/g)(Hakk et al. 2006). Preferential distribution and accumulation in adipose also has been found for hexa-brominated BDEs in studies with rats or mice exposed orally to 14C-BDE 153 (Sanders et al. 2006b) and 14C-BDE 154 (Hakk et al. 2009). In male F344 rats given single doses of 1 µmol/kg 14C-BDE 153 (~96% pure) in corn oil, levels of radioactivity remaining in tissues 24 hours after dose administration had the following order of percent of administered dose: adipose (16.7%) > muscle (13.4%) > skin (8.3%) > liver (5.8%) > blood (0.9%) > kidney (0.4%) > lung = brain (0.2% each) (Sanders et al. 2006b). Similar orders of tissueconcentrations were observed in similarly exposed female F344 rats and male and female B6C3F1 mice PBDEs 2593. HEALTH EFFECTS(Sanders et al. 2006b). During repeated daily applications of 1-µmol/kg 14C-BDE 153 doses to male rats,supralinear increases in concentrations of radioactivity (nmol 14C equivalents/g wet tissue) were observedin adipose and skin (e.g., average concentrations in adipose after 1, 3, and 10 consecutive doses were 1.57, 9.21, and 30.38 nmol/g, respectively) (Sanders et al. 2006b). This observation suggests that adiposeand skin serve as sinks for radioactivity initially distributed to other tissues. In male Sprague-Dawley rats sampled 72 hours after administration of single 12.3 µmol/kg doses of 14C-BDE 154 (>95% pure) inpeanut oil, levels of radioactivity remaining in tissues had the following order (percent of administered dose): carcass (24.3%) > gastrointestinal tract (3.77%) > adipose (1.82%) > liver (0.68) > thymus ~ testis ~ lung ~ adrenal ~ kidney ~ plasma ~ heart ~ spleen (≤0.1% each) (Hakk et al. 2009). Expression of thedata on a concentration basis (nmol 14C equivalents/g wet tissue weight) showed the highestconcentrations in lipid rich tissues: adrenals (29.56 nmol/g), adipose (21.79 nmol/g), skin (7.55 nmol/g), gastrointestinal tract (4.58 nmol/g), and thymus (4.48 nmol/g). Concentrations in carcass, lung, and liverwere intermediate (3.58, 2.26, and 1.80 nmol/g), and concentrations in testis, kidney, heart, spleen, and muscle ranged from 1.01 to 0.45 nmol/g. The concentration in plasma was very low (0.04 nmol/g) (Hakket al. 2009). Maternal transfer of decaBDE (BDE 209) has been demonstrated in female Sprague-Dawley rats givenoral 5 µmol/kg/day doses of nonlabeled BDE 209 in peanut oil from GD 7 to PPD 4 (Cai et al. 2011). BDE 209 concentrations in blood of dams increased with duration of exposure: mean concentrations (standard error [SE]) were 358.17 (210.49) and 701.17 (63.43) µg/g lipid weight on GD 15 and PPD 4, respectively (Cai et al. 2011). Whole-body concentrations of BDE 209 in fetuses and neonates also increased with duration: mean BDE 209 concentrations (and SE) were 20.53 (7.9), 28.95 (3.57), and 45.04 (10.23) µg/g lipid on GD 15, GD 21, and PPD 4, respectively (Cai et al. 2011). Nona-brominated congeners (BDE 208, BDE 207, and BDE 206) were detected in dam blood samples and whole-bodies offetuses and neonates at lower concentrations than BDE 209 and higher concentrations than octabrominated BDEs (BDE 196, 197/204, 198/203). In a related study by the same group of investigators,female Sprague-Dawley rats were exposed to 5 µmol/kg/day doses of nonlabeled BDE 209 from GD 7 to PPD 8 and from PPD 1 to 8 (Zhang et al. 2011). Whole-body BDE 209 concentrations in pups exposed during gestation and lactation were about 2 times greater than concentrations of pup exposed only duringlactation, suggesting that BDE 209 exposure and accumulation can occur during gestational and nursingperiods (Zhang et al. 2011).In another study with female Sprague-Dawley rats given 1, 10, 100, 300 or 1,000 mg/kg/day ofnonlabeled BDE 209 in corn oil from GD 7 to PPD 4, measurements of BDE 209 concentrations in blood PBDEs 2603. HEALTH EFFECTSsamples of dams and offspring gave no clear evidence of maternal transfer, but whole-bodyconcentrations in offspring were not measured (Biesemeier et al. 2010). BDE 209 concentrations in blood (reported as ng/mL blood) collected on GD 20 were generally higher in exposed dams than in exposed fetuses, but mean concentrations of dams and fetuses in the 100-, 300-, and 1,000-mg/kg groupsdid not increase with increasing dose. Mean BDE 209 concentrations in maternal or offspring blood collected on PPD 4 also did not increase with increasing dose levels, but mean BDE 209 concentrations in exposed groups of PPD 4 pups were mostly higher than mean concentrations in dams from the sameexposed groups (Biesemeier et al. 2010). Further evidence for different tissue distributions of decaBDE and lower-brominated congeners comes from a study in which neonatal NMRI mice were given single 0.7 mg/kg doses of 14C-decaBDE or0.8 mg/kg doses of 14C-BDE 99 in a 20% fat emulsion to simulate milk on PND 3, 10, or 19 (Eriksson etal. 2002b; Viberg et al. 2003a). Neonatal mice exposed to 14C-BDE 209 on PND 3, 10, or 19 had about0.48, 0.40, and 0.06% of the total administered radioactivity in the brain, 24 hours after doseadministration (Viberg et al. 2003a). Seven days after exposure, radioactivity in the brain had increasedapproximately 2-fold in mice exposed on PND 3 or 10 (to 0.74 and 1.05% of the administered dose), butremained the same as previously measured in mice exposed on PND 19 (0.06% of administered dose). Mice exposed to 14C-BDE 99 on PND 3, 10, or 19 had about 0.37, 0.51, and 0.51% of the administered dose in the brain, 24 hours after dose administration; 7 days after dose administration, decreased levels inbrains were seen in mice exposed on PNDs 3, 10, and 19 (about 0.18, 0.28, and 0.15% of theadministered dose) (Eriksson et al. 2002b). 3.4.2.3 Dermal Exposure No studies were located regarding distribution of PBDEs in humans or animals after controlled dermalexposure.3.4.3 MetabolismOxidative hydroxylation of PBDEs is a principal metabolic transformation that occurs in humans and laboratory animals. Hydroxylated PBDEs have been identified in samples of human biological fluids,including blood (Athanasiadou et al. 2008; Hovander et al. 2002; Lacorte and Ikonomou 2009; Qiu et al.2009; Rydén et al. 2012; Wan et al. 2009, 2010; Wang et al. 2012; Yu et al. 2010a) and breast milk(Lacorte and Ikonomou 2009). Hydroxylated PBDEs also have been identified in feces or bile oflaboratory rodents exposed to 14C-labeled tetra-, penta-, hexa- or decaBDEs, including: PBDEs 2613. HEALTH EFFECTS• rats orally exposed to BDE 47 (Marsh et al. 2006; Orn and Klasson-Wehler 1998; Sanders et al. 2006a);• female mice given single intravenous doses of BDE 47, BDE 99, BDE 100, or BDE 153 (Staskalet al. 2006b)• conventional or bile-duct cannulated male rats given oral or intravenous doses of BDE 99 (Chen et al. 2006; Hakk et al. 2002a);• conventional or bile-duct cannulated rats given oral doses of BDE 154 (Hakk et al. 2009); and• conventional or bile-duct cannulated rats given oral doses of BDE 209 (Morck et al. 2003; Riu etal. 2008). Hydroxylated metabolites have also been identified in feces and carcasses of male rats fed a commercialpentaBDE mixture (DE-71) for 21 days (Huwe et al. 2007), in plasma of rats given single intraperitonealinjections of an equimolar mixture of BDE 47, BDE 99, BDE 100, BDE 153, BDE 154, and BDE 183 (Malmberg et al. 2005), and in plasma of mice after oral and subcutaneous exposure to DE-71 for 34 days(Qiu et al. 2007). Oxidative hydroxylation of PBDEs also has been demonstrated and studied in in vitrometabolic systems with human liver microsomes or primary hepatocytes (Cheng et al. 2008; Erratico etal. 2012, 2013; Feo et al. 2013; Lupton et al. 2009; Stapleton et al. 2009) and rat liver microsomes orprimary hepatocytes (Dong et al. 2010; Erratico et al. 2011).Metabolic cleavage of the ether bond in PBDEs to form brominated phenols and debromination to formlower-brominated PBDEs are other metabolic fates of PBDEs in mammals. Evidence for cleavage of the ether bond includes the identification of:• glucuronide- and sulfate-conjugates of 2,4-dibromophenol in male rats given single oral doses of14C-BDE 47 (2,2',4,4'-tetraBDE) (Sanders et al. 2006a);• unconjugated 2,4,5-tribromophenol in feces and urine and glucuronide-, sulfate-, and glutathionyl-conjugates of 2,4,5-tribromophenol in bile and urine in conventional and bile ductcannulated male rats given single oral doses of 14C-BDE 99 (2,2',4,4',5-pentaBDE) (Chen et al. 2006); and • 2,4-dibromophenol and 2,4,5-tribromophenol as metabolites of BDE 47 and BDE 99 following invitro incubation with human liver microsomes (Erratico et al. 2013, 2012). PBDEs 2623. HEALTH EFFECTSEvidence for debromination includes the identification of:• hydroxylated triBDEs in feces of rats given single oral doses of 14C-BDE 47 (2,2',4,4'-tetraBDE)(Marsh et al. 2006);• hydroxylated tetraBDEs in feces of male rats given single oral doses of 14C-BDE 99(2,2',4,4',5-pentaBDE) (Hakk et al. 2002a);• mono- and di-hydroxylated tetraBDEs in feces of rats given single oral doses of 14C-BDE 100(2,2',4,4',6-pentaBDE) (Hakk et al. 2006);• hydroxylated tetra- and pentaBDEs in feces of rats given single oral doses of 14C-BDE 154(2,2',4,4',5,6'-hexaBDE) (Hakk et al. 2009);• several hydroxylated BDEs with five to seven bromines per molecule in feces of rats given single doses of 14C-BDE 209 (Morck et al. 2003); and • several hydroxylated BDEs with eight or nine bromines per molecule in male rats given singledoses of 14C-BDE 209 (Sandholm et al. 2003). Information from in vivo toxicokinetic studies with rodents exposed to the best studied congeners(i.e., BDE 47, BDE 99, BDE 100, BDE 153, BDE 154, and BDE 209) is inadequate to describe detailed metabolic pathways, but is adequate to propose that cytochrome P450s are likely to be involved in theformation of hydroxylated metabolites and hydoxylated debrominated metabolites of BDE 47 (Sanders etal. 2006a), BDE 99 (Chen et al. 2006), BDE 100 and BDE 154 (Hakk et al. 2006, 2009), and BDE 209(Morck et al. 2003; Sandholm et al. 2003). Recent in vitro studies with human or rat liver microsomes orhepatocytes, and human or rat recombinant CYPs, provide more detailed information adequate forproposing metabolic pathways for BDE 47, BDE 99, and BDE 100 in humans showing CYP2B6 mediation for hydroxylation, debromination, and ether bond cleavage for BDE 47, hydroxylation andether bond cleavage for BDE 99, and hydroxylation for BDE 100 (see Figures 3-4, 3-5, and 3-6); however, no clear metabolic pathways were identified for BDE 153 or BDE 209 using these methods.• BDE 47 was metabolized, principally by CYP2B6, in human liver microsomes (Erratico et al.2013; Feo et al. 2013; Lupton et al. 2009), producing 5-OH-BDE-47, 6-OH-BDE-47, an unidentified dihydroxylated tetrabrominated BDE and 2,4-dibromophenol as major metabolites,and several minor metabolites including three other hydroxylated tetrabrominated BDEs and two hydroxylated tribrominated BDEs (Erratico et al. 2013; see Figure 3-4). The predominance ofCYP2B6 involvement was demonstrated by inhibition of the production of all metabolites by aspecific antibody to CYP2B6 (Erratico et al. 2013) and comparison of capabilities of 11 or12 recombinant human CYPs (Erratico et al. 2013; Feo et al. 2013). PBDEs 2633. HEALTH EFFECTSFigure 3-4. Structures and General Metabolic Scheme for HydroxylatedMetabolites of BDE 47 Produced by Human Liver Microsomes*Br Br O BrBr Br Br O Br OH O OHBr Br Br O BrBr Br OHO BrBr OH Br Br O O O BrBr Br Br OH O BrBr Br OH Br O BrBr Br BrOH Br OHBr 4'-OH-BDE-17 2'-OH-BDE-28 3-OH-BDE-47 4-OH-BDE-42 5-OH-BDE-47 6-OH-BDE-47 4'-OH-BDE-49 M1 M2 x+y=4 j+k=2 x+y=4 Brx Bry OHkOHj Bry OH Brx Br 2,4-DBP *M1 and M2 refer to general structures of unidentified hydroxylated and dihydroxylated tetrabrominated BDEs. Structures of other metabolites were determined with authentic chemical standards and ultra-performance liquid chromatography-mass spectrometry techniques. Bold arrows indicate majormetabolites. CYP2B6 is proposed to be involved in production of all metabolites, based on inhibition ofBDE 47 metabolism by a specific antibody to CYP2B6, and higher rates of BDE 47 metabolism in human liver microsomes incubated with specific human recombinant CYP2B6, compared with 11 other human recombinant CYPs.Reprinted (adapted) with permission from Erratico et al. (2013). Copyright 2013 American ChemicalSociety. PBDEs 2643. HEALTH EFFECTSFigure 3-5. Structures and General Metabolic Scheme for HydroxylatedMetabolites of BDE 99 Produced by Human Liver Microsomes*O BrBr Br Br Br O BrBr Br OH Br Br O O O BrBr OH Br Br Br O BrBr Br Br BrOH O BrBr Br BrOH Br Br OHBr Br O OHBr Br Br Br Br BDE-99 4-OH-BDE-90 2,4,5-TBP 5'-OH-BDE-99 M1-M3 M4 x+y=5 j+k=2 x+y=5 Brx Bry OHkOHj Bry OH Brx 2-OH-BDE-123 4'-OH-BDE-101 6'-OH-BDE-99 *M1-3 and M4 refer to general structures of unidentified hydroxylated and dihydroxylated pentabrominated BDEs. Structures of other metabolites were determined with authentic chemicalstandards and ultra-performance liquid chromatography-mass spectrometry techniques. CYP2B6 isproposed to be involved in production of all metabolites, based on inhibition of BDE 99 metabolism by aspecific antibody to CYP2B6, and higher rates of BDE 99 metabolism in human liver microsomesincubated with human recombinant CYP2B6, compared with11 other human recombinant CYPs.Source: Erratico et al. 2012 ortho-hydroxylation (NIH shift) PBDEs 2653. HEALTH EFFECTSFigure 3-6. Structures and General Metabolic Scheme for HydroxylatedMetabolites of BDE 100 Produced by Human Liver Microsomes and Human CYP2B6*Br Br OOHBr Br Br O Br Br Br Br OH O Br OH Br Br Br O Br Br Br Br Br O Br Br OH Br Br Br O Br Br Br OH Br Br Br Br Br Br OHO O Br Br Br BrOH O Br Br OH O Br Br Br Br Br para-hydroxylation (NIH shift) 3'-OH-BDE-100 5'-OH-BDE-100 4'-OH-BDE-103 4-OH-BDE-91 2'-OH-BDE-119 6'-OH-BDE-100 3-OH-BDE-100 BDE-100 meta-hydroxylation meta-hydroxylation meta-hydroxylation para-hydroxylation ortho-hydroxylation (NIH shift) para-hydroxylation (NIH shift) ortho-hydroxylation (NIH shift) BrBr Br Br Br Br OH Br 4-OH-BDE-98 6-OH-BDE-99 *Structures of 3-OH-BDE-100, 5’-OH-BDE-100, 6’OH-BDE-100, and 4’-OH-BDE-103 were determined using commercial reference standards and ultra-performance liquid chromatography-mass spectrometrytechniques. The two remaining mono-OH-pentaBDE metabolites were hypothesized using mass spectralfragmentation characteristics of derivatized OH-BDEs. Additional information based on theoretical boiling point calculations using COnductor-like Screening MOdel for Realistic Solvents (COSMO-RS) and experimental chromatographicretention times were used to identify the hypothesized metabolites as2′-hydroxy-2,3′,4,4′,6-pentabromodiphenyl ether (2′-OHBDE-119) and 4-hydroxy-2,2′,4′,5,6-pentabromodiphenyl ether (4-OH-BDE-91), respectively. CYP2B6 is proposed to be involved in production ofall metabolites, based on inhibition of BDE 99 metabolism by a specific antibody to CYP2B6, and higherrates of BDE 100 metabolism in human liver microsomes incubated with human recombinant CYP2B6,compared with nine other human recombinant CYPs.Source: Gross et al. 2015 PBDEs 2663. HEALTH EFFECTS• BDE 47 was metabolized by liver microsomes from phenobarbital (PB)-induced and dexamethasone (DEX)-induced rats, producing five hydroxylated tetrabrominated BDEs with PB-induction and two hydroxylated tetrabrominated BDEs with DEX induction (Erratico et al. 2011). The major metabolites identified were 4'-OH-BDE-49 and 3-OH-BDE-47 (with PB or DEXinduction), and 4-OH-BDE-42 (PB induction only). No evidence for debromination or etherbond cleavage was found. These data and a comparison of metabolic capabilities of14 recombinant rat CYP enzymes indicate that rat liver metabolism of BDE 47 involves CYP1A1, CYP2A2, and CYP3A1 (Erratico et al. 2011), • BDE 99 was metabolized, principally by CYP2B6, in human liver microsomes (Erratico et al. 2012; Lupton et al. 2009; Stapleton et al. 2009), producing 2,4,5-tribromophenol, 5'-OH-BDE-99 and 4'-OH-BDE-101 as major metabolites and seven minor hydroxylated pentaBDEs (4-OHBDE-90, 6'-OH-BDE-99, and 2-OH-BDE-123, three unidentified monohydroxy pentabrominated BDEs, and one dihydroxy pentabrominated BDE) (Erratico et al. 2012; see Figure 3-5). No evidence for debromination of BDE 99 was found. CYP2B6 was the only CYP among a panel of12 human recombinant CYPs showing metabolic activity with BDE 99, and a specific antibody to CYP2B6 inhibited the production of all BDE 99 metabolites by human liver microsomes(Erratico et al. 2012).• BDE 99 was metabolized to hydroxylated metabolites in primary rat hepatocytes (Dong et al. 2010) and in liver microsomes from DEX- and PB-induced rats (Erratico et al. 2011). Livermicrosomes from DEX- and PB-induced rats produced 4-OH-BDE-99 as the major metabolite and lesser amounts (in decreasing order) of 5'-OH-BDE-99, 6'-OH-BDE-99, 2,4,5-tribromophenol, 4'-OH-BDE-101, and 2-OH-BDE-123. No evidence for debromination was found. These data and a comparison of metabolic capabilities of 14 recombinant rat CYP enzymes indicated that rat liver metabolism of BDE 99 involves CYP1A1, CYP2A2, CYP2B1, and CYP3A1 (Erratico et al. 2011).• BDE 100 was metabolized by recombinant human P450s and pooled human livermicrosomes (Gross et al. 2015). As with BDE 47 and BDE 99, human CYP2B6 was found to be the predominant enzyme responsible for nearly all formation of six mono-OHpentaBDE and two di-OH-pentaBDE metabolites. Four metabolites were identified as 3-hydroxy-2,2′,4,4′,6-pentabromodiphenyl ether (3-OH-BDE-100), 5′-hydroxy2,2′,4,4′,6-pentabromodiphenyl ether (5′-OHBDE-100), 6′-hydroxy-2,2′,4,4′,6-pentabromodiphenyl ether (6′-OH-BDE-100), and 4′-hydroxy-2,2′,4,5′,6-pentabromodiphenylether (4′-OH-BDE-103) through use of reference standards (see Figure 3-6). The two remaining mono-OH-pentaBDE metabolites were hypothesized using mass spectralfragmentation characteristics of derivatized OH-BDEs, which allowed prediction of an ortho-OH-pentaBDE and a para-OH-pentaBDE positional isomer. Additional information based on theoretical boiling point calculations using COnductor-like Screening MOdel forRealistic Solvents (COSMO-RS) and experimental chromatographic retention times wereused to identify the hypothesized metabolites as 2′-hydroxy-2,3′,4,4′,6-pentabromodiphenyl ether (2′-OHBDE-119) and 4-hydroxy-2,2′,4′,5,6-pentabromodiphenylether (4-OH-BDE-91), respectively (Simpson et al. 2015). Kinetic studies of BDE 100 metabolism using P450 2B6 and HLMs revealed Km values ranging from 4.9 to 7.0 μMand 6−10 μM, respectively, suggesting a high affinity toward the formation of OH-BDEs.Compared to the metabolism of BDE 47 and BDE 99 reported in previous studies,BDE 100 appears to be more slowly metabolized by P450s due to the presence of a third ortho-substituted bromine atom. PBDEs 2673. HEALTH EFFECTS• BDE 153 was not metabolized by human liver microsomes under conditions that produced hydroxylated metabolites from BDE 47 (a dihydroxylated BDE 47 and 2,4-dibromophenol), and BDE 99 (2,4,5- tribromophenol, a dihydroxlated BDE 99, and 1,3-dibromo benzene) (Lupton etal. 2009).• BDE 209 was not metabolized by human primary liver hepatocytes under conditions thatproduced 2,3,5-tribromophenol, two hydroxylated pentabrominated BDE metabolites and an unidentified hydroxylated tetrabrominated metabolite from BDE 99 (Stapleton et al. 2009). In summary, Feo et al. (2013) and Gross et al. (2015) characterized the in vitro metabolism of BDE 47 and BDE 100 by pooled human liver microsomes and recombinant human CYPs, identifying a number ofhydroxylated BDE metabolites; however, no brominated phenols were detected by the methods utilized by these investigators. In other studies, CYP2B6-mediated metabolism of BDE 47 and BDE 99 producesmultiple hydroxylated metabolites via hydroxylation and ether bond cleavage, based on in vitro studieswith human liver microsomes or hepatocytes and human recombinant CYPs (Erratico et al. 2012, 2013). The major metabolites of BDE 47 and BDE 99 formed by human liver microsomes were not the same as those identified using rat liver microsomes (Erratico et al. 2013, 2012, 2011). It is important to note that all studies consistently identified CYP2B6 as the primary human CYP responsible for the formation ofhydroxylated metabolites of BDE 47, BDE 99, and BDE 100 (Erratico et al. 2012, 2013; Feo et al. 2013;Gross et al. 2015), while different classes of CYP enzymes appear to be involved in in vitro rat livermetabolism of BDE 47 and BDE 99: CYP1A1, CYP2A2, and CYP3A1 for BDE 47 and CYP1A1, CYP2A2, CYP2B1, and CYP3A1 for BDE 99 (Erratico et al. 2011). Production of hydroxylated metabolites of BDE 153 (Lupton et al. 2009) and BDE 209 (Stapleton et al. 2009) has not been demonstrated with human liver microsomes or hepatocytes, respectively. It is uncertain if these latterfindings are reflective of a limited in vivo capacity of humans to metabolize these BDE congeners orbecause the proper in vitro conditions for metabolizing these congeners were not provided. Currently, studies of metabolism of BDE 47 and BDE 99 with in vitro human and rat systems have found evidence of metabolic oxidative debromination only with BDE 47 in human liver microsomes. In contrast, anumber of in vivo studies have found evidence for oxidative debromination in feces collected from ratsexposed to BDE 47 (Marsh et al. 2006), BDE 99 (Hakk et al. 2002a), BDE 100 (Hakk et al. 2006), BDE 154 (Hakk et al. 2009), and BDE 209 (Morck et al. 2003; Sandholm et al. 2003).3.4.4 Elimination and Excretion 3.4.4.1 Inhalation Exposure No studies were located regarding excretion of PBDEs in humans or animals after inhalation exposure. PBDEs 2683. HEALTH EFFECTS3.4.4.2 Oral ExposureOral Exposure Elimination Overview. Apparent half-lives of PBDE congeners in blood of PBDE-exposed workers during non-exposed vacation periods ranged from 15 days for BDE 209, 18–39 days fornonabrominated congeners, and 37–94 days for octabrominated congeners. The detection of PBDEs in human breast milk samples indicates that elimination via milk is an elimination route for women, butseveral studies examining PBDE concentrations during lactation do not provide a clear account of thedegree to which PBDEs are cleared from the body during lactation. Results from animal studies given single oral doses of 14C-labeled PBDE congeners or PBDE mixtures indicate that biliary excretion into thefeces is the principal route of elimination in rats, and that the urine and feces are principal routes ofelimination of orally absorbed PBDEs in mice.No studies were located regarding excretion of PBDEs in humans after controlled oral exposure.Apparent half-lives of several PBDEs in blood were estimated for eight PBDE-exposed workers (fourelectronics dismantlers and four workers in a factory making flame-retarded rubber) from measurements of hepta-, octa-, nona- and decaBDE concentrations in blood collected during 28–29-day vacation periodswithout occupational exposure (Thuresson et al. 2006). After fitting the data to a single-phase exponential model, calculated apparent half-lives were: 15 days for BDE 209; 28, 39, and 18 days fornonabrominated congeners BDE 208, BDE 207, and BDE 206; 37, 72, 85, and 91 days for BDE 203 and three other octabrominated congeners of uncertain chemical structure; and 94 days for BDE 183.The detection of PBDEs in human breast milk samples suggests that breast milk represents an eliminationroute of absorbed PBDEs in women (see Jakobsson et al. 2012 and Frederiksen et al. 2009 for reviews ofPBDE levels in breast milk). Several studies have examined changes in PBDE concentrations (and otherpersistent lipophilic chemicals) in breast milk during lactation, but the results do not provide a clearaccount of the degree to which PBDEs are cleared from the body during breast feeding (Hooper et al. 2007; Jakobsson et al. 2012; LaKind et al. 2009; Thomsen et al. 2010).Studies with rats given single oral doses of 14C-labeled PBDE congeners or PBDE mixtures indicate thatingested PBDEs are principally excreted in the feces with <2% of administered radioactivity excreted in the urine within 3 days of dose administration. This pattern has been observed in male and female rats exposed to BDE 47 (Orn and Klasson-Wehler 1998; Sanders et al. 2006a), male and female rats exposed PBDEs 2693. HEALTH EFFECTSto BDE 99 (Chen et al. 2006; Hakk et al. 2002a), male rats exposed to BDE 100 (Hakk et al. 2006), maleand female rats exposed to BDE 153 (Sanders et al. 2006), male rats exposed to BDE 154 (Hakk et al.2009), pregnant female and male rats exposed to BDE 209 (Morck et al. 2003; Riu et al. 2008), and male and female rats exposed to a mixture with 77.4% 14C-BDE 209, 21.8% nonabrominated BDE, and 0.8%octabrominated (Norris et al. 1973, 1975b). Studies with bile duct-cannulated male rats indicated thatradioactivity excreted in feces of conventional rats included bile duct-delivered metabolites (i.e., absorbedmaterial) and unabsorbed compound. Cumulative biliary excretion represented about 3.9% ofadministered dose with BDE 99 (Hakk et al. 2002a), 1.7% with BDE 100 (Hakk et al. 2006), 1.3% with BDE 154, and about 10% with BDE 209 (Morck et al. 2003). A different elimination pattern has been observed in mice, especially with BDE 47. In mice given single oral doses of 14C-labeled PBDE congeners, fecal and urinary elimination were principal routes ofelimination for BDE 47 (Orn and Klasson-Wehler 1998; Sanders et al. 2006a), whereas fecal elimination appeared to be more important than urinary elimination with BDE 99 (Chen et al. 2006) and BDE 153(Sanders et al. 2006b). Male C57Bl mice excreted 20% of administered radioactivity in feces and 33% in urine within 5 days of administration of 14C-BDE 47 (Orn and Klasson-Wehler 1998). Male and female B6C3F1 mice excreted about 30 and 20% of administered 14C-BDE 47 dose in urine and about 22% and 25% in feces, within 24 hours (Sanders et al. 2006a). In B6C3F1 mice given 14C-BDE 99, males excreted7.8% in urine and 27.1% in feces within 24 hours, and females excreted 4.1% in urine and 32.4% in feces(Chen et al. 2006). Twenty-four hours after administration of 14C-BDE 153, male B6C3F1 mice excreted1% of administered dose in urine and 31.5% in feces, and females excreted 0.3% in urine and 26.0% in feces (Sanders et al. 2006b).Complementary studies with female C57BL/6J given single intravenous 1-mg/kg doses of 14C-labeledBDE 47, BDE 99, BDE 100, or BDE 153 also indicate that the degree of urinary excretion in mice iscongener specific (Staskal et al. 2006b). Cumulative percentages of administered radioactivity excretedin urine within 5 days were 40, 16, 6, and 2% for BDE 47, BDE 99, BDE 100, and BDE 153, respectively. Relatively greater amounts of parent compound were found in urine from BDE 47-exposed mice, compared with mice exposed to the other congeners. Ratios of cumulative percentage dose excreted as parent compound or metabolite in urine were 1.5, 0.5, 0.4, and 0.2 for BDE 47, BDE 99, BDE 100, and BDE 153, respectively. In feces, respective ratios of parent compound:metabolite were 0.7, 0.2, 0.2, and 0.3 for these congeners, respectively. PBDEs 2703. HEALTH EFFECTSThe mechanism for the relatively high elimination of BDE 47 in the urine of mice is unknown, but hasbeen hypothesized to involve binding to mouse major urinary protein (m-MUP) in the blood (Sanders etal. 2006a; Staskal et al. 2006b) and other membrane transporting polypeptides (Emond et al. 2013;Pacyniak et al. 2010, 2011).3.4.4.3 Dermal Exposure No studies were located regarding excretion of PBDEs in humans or animals after dermal exposure.3.4.5 Physiologically Based Pharmacokinetic (PBPK)/Pharmacodynamic (PD) Models Physiologically based pharmacokinetic (PBPK) models use mathematical descriptions of the uptake anddisposition of chemical substances to quantitatively describe the relationships among critical biologicalprocesses (Krishnan et al. 1994). PBPK models are also called biologically based tissue dosimetrymodels. PBPK models are increasingly used in risk assessments, primarily to predict the concentration ofpotentially toxic moieties of a chemical that will be delivered to any given target tissue following variouscombinations of route, dose level, and test species (Clewell and Andersen 1985). Physiologically based pharmacodynamic (PBPD) models use mathematical descriptions of the dose-response function to quantitatively describe the relationship between target tissue dose and toxic end points. PBPK/PD models refine our understanding of complex quantitative dose behaviors by helping to delineate and characterize the relationships between: (1) the external/exposure concentration and targettissue dose of the toxic moiety, and (2) the target tissue dose and observed responses (Andersen and Krishnan 1994; Andersen et al. 1987). These models are biologically and mechanistically based and can be used to extrapolate the pharmacokinetic behavior of chemical substances from high to low dose, fromroute to route, between species, and between subpopulations within a species. The biological basis ofPBPK models results in more meaningful extrapolations than those generated with the more conventionaluse of uncertainty factors.The PBPK model for a chemical substance is developed in four interconnected steps: (1) modelrepresentation, (2) model parameterization, (3) model simulation, and (4) model validation (Krishnan andAndersen 1994). In the early 1990s, validated PBPK models were developed for a number oftoxicologically important chemical substances, both volatile and nonvolatile (Krishnan and Andersen 1994; Leung 1993). PBPK models for a particular substance require estimates of the chemical substance-specific physicochemical parameters, and species-specific physiological and biological parameters. The PBDEs 2713. HEALTH EFFECTSnumerical estimates of these model parameters are incorporated within a set of differential and algebraic equations that describe the pharmacokinetic processes. Solving these differential and algebraic equations provides the predictions of tissue dose. Computers then provide process simulations based on thesesolutions. The structure and mathematical expressions used in PBPK models significantly simplify the true complexities of biological systems. If the uptake and disposition of the chemical substance(s) are adequately described, however, this simplification is desirable because data are often unavailable formany biological processes. A simplified scheme reduces the magnitude of cumulative uncertainty. The adequacy of the model is, therefore, of great importance, and model validation is essential to the use ofPBPK models in risk assessment.PBPK models improve the pharmacokinetic extrapolations used in risk assessments that identify themaximal (i.e., the safe) levels for human exposure to chemical substances (Andersen and Krishnan 1994).PBPK models provide a scientifically sound means to predict the target tissue dose of chemicals inhumans who are exposed to environmental levels (for example, levels that might occur at hazardous waste sites) based on the results of studies where doses were higher or were administered in different species.Figure 3-7 shows a conceptualized representation of a PBPK model.If PBPK models for PBDEs exist, the overall results and individual models are discussed in this section in terms of their use in risk assessment, tissue dosimetry, and dose, route, and species extrapolations.Emond et al. (2010) developed a PBPK model for BDE 47 in male and female (nonpregnant and pregnant) adult rats. The model included eight compartments: liver, brain, adipose tissue, kidney, placenta, fetus, blood, and the remaining body. The model was calibrated with tissue concentration data from adult male and maternal-fetal toxicokinetic studies. Other data sets were then used to evaluate the model’s performance. Model evaluations indicated that simulated BDE 47 tissue concentrations in adultmale, adult female, and fetal compartments were within the standard deviations of the empirical data.Emond et al. (2013) developed a PBPK model for BDE 47 in adult mice to describe the distribution ofBDE 47 in tissues and its elimination in feces and urine, and to evaluate the role of transporters in elimination of BDE 47. The structure of the model was similar to the rat model developed by Emond etal. (2010), without the gestational submodel. In addition, binding to transporters proposed to facilitate PBDEs 2723. HEALTH EFFECTSFigure 3-7. Conceptual Representation of a Physiologically BasedPharmacokinetic (PBPK) Model for a Hypothetical Chemical SubstanceInhaled chemical Exhaled chemical Lungs Liver Fat Slowly perfused tissues Richly perfused tissues Kidney Skin V E N O U S B L O O D A R T E R I A L B L O O D Vmax Km Ingestion GI Tract Feces Urine Chemicals contacting skin Note: This is a conceptual representation of a physiologically based pharmacokinetic (PBPK) model for a hypothetical chemical substance. The chemical substance is shown to be absorbed via the skin, by inhalation, or byingestion, metabolized in the liver, and excreted in the urine or by exhalation.Source: Krishnan and Andersen 1994 PBDEs 2733. HEALTH EFFECTSurinary excretion in mice were added: mouse major urinary protein (m-MUP) in blood and P-glycoprotein, a membrane transporter in brain, liver, and kidneys. The model was used to investigatethe roles that m-MUP and P-glycoprotein may play in BDE 47 elimination in mice.No reports were located on the development of PBPK models for PBDEs in humans.3.5 MECHANISMS OF ACTION 3.5.1 Pharmacokinetic MechanismsAbsorption. Information regarding oral absorption in animals is available from studies of commercialPBDE mixtures and individual 14C-labeled tetra-, penta-, hexa-, and decaBDE congeners. The mostrecent and best available estimates of oral absorption efficiencies following gavage administration inlipophilic vehicles indicate a range of 70–85% for tetraBDE (BDE 47), pentaBDE (BDE 99, BDE 100),and hexaBDE (BDE 153, BDE 154) congeners, and 10–26% for decaBDE (BDE 209) (Hakk et al. 2001, 2002b, 2006, 2009; Huwe et al. 2002b, 2007; Morck and Klasson Wehler 2001; Morck et al. 2003; Örn and Klasson-Wehler 1998; Riu et al. 2008; Sanders et al. 2006a, 2006b; Sandholm et al. 2003; Staskal etal. 2005). Underlying mechanisms for oral absorption (e.g., active transport, diffusion, protein binding)have not been described.Studies using in vitro gastrointestinal digestion models have evaluated bioaccessibility of PBDEs inenvironmentally relevant sources. In a study by Yu et al. (2010b), the bioaccessibility of lowerbrominated PBDEs in 13 types of food (fish, meat, rice, flour, and vegetables) ranged from 2.6 to 41.3%in food. Dietary fat was the most important factor affecting the bioaccessibility of PBDEs, withbioavailability increasing with increased fat content, likely due to the lipophilic nature of PBDEs (Yu et al. 2010b). Bioavailability was also increased with increasing carbohydrate content, potentially due to theformation of micelles (Yu et al. 2010b). However, bioavailability decreased with increasing protein andfiber content, potentially due to adsorption to dietary fiber and ionic strength effect of amino acids leadingto decreased partition of PBDEs in the aqueous phase (Yu et al. 2010b). Lepom et al. (2010) evaluated the bioaccessibility of PBDEs in ingested dust, which is expected to be the predominant source of human exposure in the United States (EPA 2010). In this study, the bioavailability of PBDEs in ingested dustwas <50%, with higher bioavailability for the lower-brominated PBDEs (27–42%) compared with BDE 209 (10%) (Lepom et al. 2010). A similar study by Abdallah et al. (2012) showed comparableresults for the bioaccessibility of PBDEs in ingested dust, with higher bioavailability for the lowerbrominated PBDEs (32–58%) compared with BDE 209 (14%). PBDEs 2743. HEALTH EFFECTSA dermal absorption efficiency of 62% was reported for female mice exposed to an occluded dermal dose of 1 mg/kg 14C-BDE 47) (Staskal et al. 2005). In vitro studies have evaluated diffusion of PBDEs across human, rat, and mouse skin. For 14C-BDE 47, the mean absorption efficiencies (percent of appliedradioactivity collected in receptor compartment) were 1.88% for human skin and 14.58% for rat skin (Roper et al. 2006). For 14C-decaBDE, mean absorption efficiency for mouse skin in vitro was reportedto be only 0.07–0.34% of the applied dose (Hughes et al. 2001). Underlying mechanisms for dermalabsorption (e.g., active transport, diffusion, protein binding) specific to PBDEs have not been described.Distribution. Evidence for the transfer of PBDEs from pregnant mothers to the developing fetus and forthe transfer of PBDEs from maternal blood to breast milk and then to nursing infants comes from anumber of studies of PBDE concentrations in maternal and cord serum samples and breast milk samples from groups of non-occupationally exposed women (Antignac et al. 2009, 2008; Chen et al. 2013;Kawashiro et al. 2008; Li et al. 2013b; Malarvannan et al. 2013; Mazdai et al. 2003; Meijer et al. 2008;Park et al. 2011; Qiu et al. 2009; Schecter et al. 2010; Vizcaino et al. 2011; Wan et al. 2010). In general, the tetra- and penta-brominated PBDEs have been the predominant congeners detected in maternal andcord serum samples and breast milk samples, but some recent studies analyzing a wider range of PBDEcongeners have found evidence for distribution of hepta-, octa-, or decaBDEs into cord serum and breastmilk. Frederiksen et al. (2010) utilized a human ex vivo placenta perfusion system to study the kinetics ofplacental transfer of BDE 47, BDE 99, and BDE 209 to the fetus. Placentas were perfused for 4 hourswith 1 ng/mL concentrations of the nonlabeled congeners. Samples of maternal and fetal compartmentswere taken during the perfusion, and concentrations in samples were determined by GC/MS. Placentaltransfer of BDE 47 and BDE 99 was demonstrated, and transfer of BDE 47 was faster and more extensivethan BDE 99. Transfer of BDE 209 across the placenta to the fetal compartment was not demonstrated with the detection limits of the techniques employed. Frederiksen et al. (2010) proposed that futureperfusion studies with BDE 209 should use 14C-labeled BDE 209 to increase sensitivity.Tissue distribution studies in animals orally exposed to 14C-labeled BDE 47, BDE 99, BDE 100, BDE 153, BDE 154, and BDE 209 indicate that decaBDE is distributed among tissues somewhatdifferently than tetra-, penta- and hexaBDEs. While lower-brominated BDE congeners, followingabsorption and an initial wide distribution, are preferentially accumulated in adipose tissues, absorbeddecaBDE is less readily distributed to adipose tissues and appears to preferentially distribute to highlyperfused tissues (Chen et al. 2006; El Dareer et al. 1987; Eriksson et al. 2002b; Hakk et al. 2002a, 2006;Morck and Klasson-Wehler 2001; Morck et al. 2003; Norris et al. 1975a; NTP 1986; Örn and Klasson PBDEs 2753. HEALTH EFFECTSWehler 1998; Riu et al. 2008; Sanders et al. 2006a, 2006b; Staskal et al. 2005, 2006a; Viberg et al.2003a). Although less likely to partition to adipose tissues, decaBDE was still found in low quantities in adipose tissues in these studies, and has been shown to transfer from dams to fetuses and neonates fromexposure during gestational and nursing periods (Cai et al. 2011; Zhang et al. 2011). Metabolism. Oxidative hydroxylation of PBDEs is a principal metabolic transformation that is thought to occur in humans and laboratory animals. Hydroxylated PBDEs have been identified in samples of human biological fluids including blood (Athanasiadou et al. 2008; Hovander et al. 2002; Lacorte and Ikonomou 2009; Qiu et al. 2009; Rydén et al. 2012; Wan et al. 2009, 2010; Wang et al. 2012; Yu et al. 2010a) and breast milk (Lacorte and Ikonomou 2009). Hydroxylated PBDEs also have been identified in feces orbile of laboratory rodents exposed to 14C-labeled tetra-, penta-, hexa- or decaBDEs (Chen et al. 2006;Hakk et al. 2002a, 2009; Marsh et al. 2006; Morck et al. 2003; Orn and Klasson-Wehler 1998; Riu et al. 2008; Sanders et al. 2006a; Staskal et al. 2006b). Oxidative hydroxylation of PBDEs also has been demonstrated and studied in in vitro metabolic systems with human liver microsomes or primaryhepatocytes (Cheng et al. 2008; Erratico et al. 2012, 2013; Feo et al. 2013; Lupton et al. 2009; Stapleton et al. 2009) and rat liver microsomes or primary hepatocytes (Dong et al. 2010; Erratico et al. 2011). Metabolic cleavage of the ether bond in PBDEs to form brominated phenols and debromination to formlower-brominated PBDEs are other metabolic fate processes for PBDEs in mammals (Chen et al. 2006;Erratico et al. 2013, 2012; Hakk et al. 2002a, 2006, 2009; Marsh et al. 2006; Morck et al. 2003; Sanders et al. 2006a; Sandholm et al. 2003).Information from in vivo toxicokinetic studies with rodents exposed to the best studied congeners(i.e., BDE 47, BDE 99, BDE 100, BDE 153, BDE 154, and BDE 209) is inadequate to describe detailed metabolic pathways, but is adequate to propose that cytochrome P450s are likely to be involved in theformation of hydroxylated metabolites and hydoxylated debrominated metabolites. In vitro studies withhuman liver microsomes or hepatocytes and human recombinant CYPs indicate that CYP2B6-mediatedmetabolism of BDE 47, BDE 99, and BDE 100 produced multiple metabolites via hydroxylation (Erratico et al. 2012, 2013; Feo et al. 2013; Gross et al. 2015) and ether bond cleavage (Erratico et al. 2012, 2013). The major metabolites of BDE 47 and BDE 99 formed by human liver microsomes were not the same as those identified using rat liver microsomes (Erratico et al. 2011, 2012, 2013). Different sets of CYPenzymes appear to be involved in in vitro rat liver metabolism of BDE 47 and BDE 99: CYP1A1, CYP2A2, and CYP3A1 for BDE 47 and CYP1A1, CYP2A2, CYP2B1, and CYP3A1 for BDE 99 (Erratico et al. 2011). Production of hydroxylated metabolites of BDE 153 (Lupton et al. 2009) and BDE209 (Stapleton et al. 2009) has not been demonstrated with human liver microsomes or hepatocytes. It is PBDEs 2763. HEALTH EFFECTSuncertain if these latter findings are reflective of a limited in vivo capacity of humans to metabolize these BDE congeners or because the proper in vitro conditions for metabolizing these congeners were notprovided. While rat studies provide evidence for metabolic oxidative debromination of BDE 47, BDE 99,BDE 100, BDE 154, and BDE 209, studies with human liver microsomes only found evidence supportingthe oxidative debromination of BDE 47.Elimination. Apparent half-lives of PBDE congeners in blood of PBDE-exposed workers during non-exposed vacation periods ranged from 15 days for BDE 209, 18–39 days for nonabrominated congeners, and 37–94 days for octabrominated congeners (Thuresson et al. 2006). The detection of PBDEs in human breast milk samples indicates that elimination via milk is an elimination route for women, but severalstudies examining PBDE concentrations during lactation do not provide a clear account of the degree to which PBDEs are cleared from the body during lactation (Hooper et al. 2007; Jakobsson et al. 2012;LaKind et al. 2009; Thomsen et al. 2010). Results from animal studies given single oral doses of14C-labeled PBDE congeners or PBDE mixtures indicate that biliary excretion into the feces is the principal route of elimination in rats (Chen et al. 2006; Hakk et al. 2002a, 2006; Morck et al. 2003; Norriset al. 1973, 1975b; Orn and Klasson-Wehler 1998; Riu et al. 2008; Sanders et al. 2006a), and that theurine and feces are principal routes of elimination of orally absorbed PBDEs in mice (Chen et al. 2006;Orn and Klasson-Wehler 1998; Sanders et al. 2006a). In mice, the importance of urinary excretion iscongener-specific, with BDE 47 showing the greatest ratio of cumulative percentage dose excreted as parent compound or metabolite (Staskal et al. 2006b). The mechanism for the relatively high eliminationof BDE 47 in the urine of mice is unknown, but has been hypothesized to involve binding to mouse majorurinary protein (m-MUP) in the blood (Sanders et al. 2006a; Staskal et al. 2006b) and other membranetransporting polypeptides (Emond et al. 2013; Pacyniak et al. 2010, 2011).3.5.2 Mechanisms of ToxicityOverview. As summarized in Section 2.2 (Summary of Health Effects) and detailed in Chapter 3 (Health Effects), the main targets of concern following PBDE exposure in humans are the developing nervous and reproductive systems, the developing and mature endocrine system, the liver, and the male reproductive system. Other potential targets are the female reproductive system, the adult nervous system, and the developing and adult immune system; however, evidence for these end points is limited. Numerousstudies have been conducted to identify potential mechanisms of toxicity for PBDE exposure. These studies include evaluations of general mechanisms (e.g., hepatic enzyme induction, AhR-mediatedeffects) as well as target-specific mechanisms. For specific targets, the majority of mechanistic studies PBDEs 2773. HEALTH EFFECTShave focused on endocrine disruption and neurological effects; however, definitive mechanismsunderlying these effects have not been elucidated. For other effects, including reproductive toxicity,immunotoxicity, and hepatotoxicity, only limited mechanistic data are available. Mechanistic data relevant to toxic effects of PBDEs are reviewed below.General Mechanisms of Toxicity. PBDEs share some toxicological properties with other structurallysimilar polyhalogenated aromatic compounds, particularly PBBs, PCBs, PCDDs, and PCDFs (ATSDR1994, 1998, 2000). Although these chemicals are structurally similar in two dimensions, PBDEs (and polychlorinated diphenyl ethers or PCDEs) differ from the other classes on a three-dimensional basis. In particular, the oxygen bridge of the ether linkage in the diphenyl ether molecule increases the distance between the biphenyl rings, introduces a 120° bend in the alignment of the biphenyl rings, and serves as abarrier to rotation that inhibits the two aromatic rings from assuming a coplanar configuration (Hardy2002a; Howie et al. 1990). Furthermore, halogen substitution of the ortho positions in the diphenylethermolecule, as occurs for some congeners, pushes the aromatic rings to be orthoganol to each other (i.e., offset by 90°) (Hardy 2002a). Because the toxicity of TCDD and related compounds is related to their ability to assume a coplanar configuration for binding to AhR, this suggests that PBDEs are unlikely to display similar toxic potency (Hardy 2002a). Assays conducted by Chen et al. (2001) to compare AhRbinding affinity and EROD activity of PBDE congeners and mixtures found that activities were very lowrelative to TCDD and related compounds, but also that, unlike for PCBs, AhR binding affinity was notcorrelated with planarization energies of the congeners (the calculated energy needed to force coplanarity of the PBDE molecule). The researchers speculated that the large size of the bromine atoms may distortthe AhR binding site so that coplanar configuration is not required. However, even in studies ofchlorinated analogs (PCDEs), it was found that increasing ortho substitution is less effective in decreasingthe activity of these congeners than for PCBs (Howie et al. 1990). The authors attributed this differenceto the increased bond length between the phenyl rings in PCDEs relative to PCBs.In other words, the ether bridge makes PBDEs more non-coplanar in nature, which reduces AhR binding affinity relative to similar compounds, but also less sensitive to the influence of ortho substitutions thatinhibit AhR binding of PBBs or PCBs. This has implications not only for dioxin-type toxicities, whichare mediated by the AhR pathway, but also for non-dioxin-type effects. For example, Chen et al. (2001)found that the induction of CYP1A1 by PBDEs is AhR-mediated, as it is for numerous organochlorines, even though PBDEs do not readily adopt the coplanar conformation usually considered characteristic ofAhR ligands. Structure-activity relationships have been incompletely elucidated for non-dioxin-likeeffects of PBDEs such as neurotoxicity. However, based on limited available data, it can be speculated PBDEs 2783. HEALTH EFFECTSthat di-ortho-substituted PBDEs might follow the neurotoxic potency of ortho-PCBs (Eriksson et al. 2002b; Kodavanti and Derr-Yellin 2002; Mariussen and Fonnum 2002, 2003).There are also geometrical differences in PCBs, PBBs, and PBDEs due to the higher atomic weight andconsiderably larger molecular volume of bromine compared to chorine (Hardy 2000a, 2002a). Thesedifferences contribute to dissimilar physical/chemical properties that can influence the relativebioavailability, absorption, tissue accumulation, receptor interactions, and toxicities of the chemicals. Forexample, a comparison of a series of isosteric 3,3',4,4'-tetrahalobiphenyls in rats showed that relative toxicity (growth rate and thymic atrophy), AhR binding affinity, and aryl hydrocarbon hydroxylase(AHH) and EROD induction potencies increased with increasing bromine substitution (Andres et al. 1983). Possible explanations for this effect included the increased polarizability of bromine versuschlorine and differences in the electronic, hydrophobic, and hydrogen bonding characteristics of bromineand chlorine (Andres et al. 1983). The enzyme induction properties of PBDEs have been studied to a lesser extent than the enzyme induction properties of other structurally similar chemicals. Existing information suggests that PBDEs can be classified as mixed-type inducers of hepatic microsomal monooxygenases, although the mixed induction properties of the commercial mixtures are likely due to contamination with polybrominatedp-dibenzodioxins (PBDDs) and polybrominated dibenzofurans (PBDFs) (Darnerud et al. 2001; de Wit2002; Hardy 2002b). Few studies have examined the structure-induction relationships for PBDEs. Chen et al. (2001) examined the ability of 12 PBDE congeners and 3 commercial mixtures to induce ERODactivity in chick and rat hepatocytes, in liver cell lines from rainbow trout, rat, and human, and in ahuman intestinal cell line. The number of bromine substitutions in the congeners tested ranged from 3 to 7. In all cell types, BDE 77, 2,2',4,4',6-pentaBDE, BDE 66, and BDE 126 were the strongest inducers. BDE 153 and 2,2',3,4,4',5',6-heptaBDE were weak inducers in all cell types, whereas BDE 66 and BDE 85 were very weak inducers in rat hepatocytes and inactive in the other cells. BDE 47 and 2,2'4,4',5-pentaBDE, which are prominent in the environment, were not inducers in any cell line, and neither were BDE 28, 2,2',4,4',5,6'-hexaBDE, or the penta-, octa-, or decaBDE mixtures. For thosecongeners that had measurable EROD induction activity, their relative potencies were 10-3–10-6 that of2,3,7,8-TCDD. In general, the EROD induction activity paralleled the strength of the AhR binding with the notable exception of BDE 85, which despite its relatively strong AhR binding affinity (see above),showed no evidence of activating the AhR to its DRE binding form and was only a weak EROD inducer. PBDEs 2793. HEALTH EFFECTSAs discussed in the introduction to this section, bromination at the ortho position does not appear tosignificantly change the biological effects of PBDE molecules. Structure-activity studies have shown thatsome PBDE congeners can bind to the AhR, although binding affinities and induction of AhR-mediatedresponses are very weak or negligible, particularly for commercial PBDE mixtures and environmentallyrelevant congeners.For example, Meerts et al. (1998) indirectly examined the AhR-mediated (dioxin-like) properties of17 PBDE congeners in a recombinant H4II rat hepatoma cell line showing AhR-mediated expression of aluciferase reporter gene. The tested congeners varied from dibromo- to heptabromo- substitutedcompounds, and with the exception of BDE 15 and BDE 77, all had at least one ortho substitution. Seven of the congeners showed luciferase expression, indicating their ability to activate the AhR. The onlydiscernable pattern of receptor activation that appeared to emerge from these results was that greaterreceptor activation was obtained with the penta- and hexaBDEs than with tri- and tetraBDEs.Another study also examined the AhR induction potency of PBDE congeners using the in vitro luciferaseassay with H4IIE-luc recombinant rat hepatoma cells (Villeneuve et al. 2002). Only 1 of 10 testedcongeners (BDE 126) induced a significant response for AhR-mediated gene expression in the H4IIE-luccells, but the magnitude of induction was 87% less compared with the response induced by 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD). With the exception of BDE 105, which induced a responseof 1.7% of the TCDD maximum, no other congener, including the environmentally prominent congeners,BDE 47, BDE 99, and BDE 153, yielded a response greater than 1% of TCDD. Overall, the tested PBDEcongeners were at least 200,000 times less potent than TCDD for inducing AhR-mediated gene expression in this test system. Using the same test system (H4IIE-luc recombinant rat hepatoma cells),Behnisch et al. (2003) reported AhR activities 5–6 orders of magnitude lower than TCDD for BDE 77, BDE 105, BDE 126, BDE 119, BDE 190, and BDE 209; BDE 47, BDE 66, BDE 85, BDE 99, BDE 100, BDE 153, and BDE 183 were inactive. Similarly, Hamers et al. (2006) reported AhR activities 6 orders of magnitude lower than TCDD for 7/19 PBDE congeners tested (BDE 38, BDE 49, BDE 79, BDE 99, BDE 181, BDE 183, and BDE 190), as well as the hydroxylated metabolite, 6OH-BDE-47. In anotherstudy, several hydroxylated and methoxylated PBDEs (19 out of 34 tested) were also shown to activate the AhR receptor in the H4IIE-luc assay, with potencies 4–12 orders of magnitude less than TCDD (Su etal. 2012).Chen et al. (2001) studied the affinities of a series of 18 PBDE congeners and 3 commercial PBDEmixtures for rat hepatic AhR by using competitive AhR-ligand and EROD induction assays. The analysis PBDEs 2803. HEALTH EFFECTSshowed that both the congeners and octa- and pentaBDE commercial mixtures had binding affinities of10-2–10-5 times that of 2,3,7,8-TCDD. The congener with the highest affinity among the tested congenerswas BDE 85, although its relative binding affinity was only 2% that of 2,3,7,8-TCDD. No bindingactivity could be determined for the decaBDE mixture. In contrast with PCBs, the binding affinities did not appear to relate to the planarity of the molecule, which according to Chen et al. (2001), was possiblydue to the fact that the large size of bromine atoms expands the receptor binding site. The dioxin-likeactivity of the PBDE congeners and commercial mixtures was subsequently more completelycharacterized, by determining whether they act as AhR agonists or antagonists at sequential stages of the AhR signal transduction pathway leading to CYP1A1 expression in rat hepatocytes (Chen and Bunce 2001). BDE 77, BDE 119, and BDE 126 were moderately active towards dioxin-response element (DRE)binding and induced responses of both CYP1A1 mRNA and CYP1A1 protein analogous to the maximal response of TCDD, although at concentrations 3–5 orders of magnitude greater than TCDD. These congeners showed additive behavior towards DRE binding with TCDD (i.e., an increased responsecompared to TCDD alone), whereas most of the other congeners antagonized the action of TCDD. BDE100, BDE 153, and BDE 183 were very weak activators of DRE binding, and other congeners and thethree commercial BDE mixtures were inactive. In particular, the environmentally prominent congenersBDE 47 and BDE 99 were among the least active with respect to dioxin-like behavior (i.e., were inactiveat all stages of signal transduction), and the commercial pentaBDE mixture had negligible ERODinduction activity. The PBDE congeners that bound most strongly to the AhR were also the strongestinducers of CYP1A1 mRNA and CYP1A1 protein, indicating that the induction of CYP1A1 was AhRmediated. Considering all of the end points evaluated in the Chen et al. (2001) and Chen and Bunce(2001) studies, it was concluded that the relative induction potencies (REPs) of the most active PBDEs toward CYP1A1 are ≈10-4 that of TCDD, which is similar to some mono-ortho-PCBs and two orders of magnitude less than those of coplanar PCBs, but the REPs for the environmentally prominent congenersare essentially zero. Consistent with these findings, Behnisch et al. (2003) also reported REPs of 10-4– 10-6 that of TCDD for BDE 25, BDE 77, BDE 100, BDE 126, and BDE 183 in the H4IIE-EROD bioassay, while BDE 154, BDE 99, BDE 47, and BDE 28 were inactive.Mechanism of Endocrine Disruption. PBDE-induced endocrine disruption is likely to involve multiplemechanisms, including altered synthesis/clearance, transport, and/or receptor binding of endocrinehormones.The apparent lack of effect of PBDEs on serum TSH suggests that direct effects on the thyroid leading to inhibition of T4 synthesis are unlikely. However, Wu et al. (2016) present in vitro data that BDE 47 PBDEs 2813. HEALTH EFFECTSinhibits iodide uptake in rat thyroid follicular FRTL-5 cells via non-competitive inhibition of thesodium/iodide symporter (NIS). Additionally, mRNA expression levels of thyroid peroxidase (TPO) wasdecreased by BDE 47. Together, NIS and TPO are considered critical molecules in thyroid hormonesynthesis, and are targets of many thyroid disruptors (e.g., perchlorate and thiocyanate) (Wu et al. 2016). No additional data regarding the potential for PBDEs to interfere with thyroid hormone synthesis werelocated.PBDEs are hepatic microsomal enzyme inducers, but there is little evidence that increased enzyme activity leads to greater clearance of thyroid hormones. The induction of hepatic UDPGT by PBDEs has been demonstrated in several studies (Fowles et al. 1994; Hallgren and Darnerud 2002; Hallgren et al.2001; Skarman et al. 2005; Stoker et al. 2004; Szabo et al. 2009; Zhou et al. 2001, 2002) and this could increase the UDPGT-catalyzed deactivation and excretion of T4 (i.e., the conjugation of T4 withglucuronic acid). An indication that increased UDPGT activity may not be the main mechanism for thereduced T4 levels is provided by Hallgren et al. (2001), who found that exposure to ≥18 mg/kg/day pentaBDE for 14 days caused serum T4 reductions in both mice and rats with no effect on UDPGTactivity in the mice, and increased UDPGT in the rats only at higher dose levels. In contrast, the decreases in serum T4 correlated with the induction of microsomal phase I enzymes (EROD and MROD). Increased microsomal enzyme activity (discussed above) could also increase the formation ofhydroxylated PBDE metabolites that can bind to T4 plasma transport proteins. This would serve toincrease the number of occupied sites on T4-binding proteins and subsequently result in decreased serumlevels of T4; however, this mechanism is not fully elucidated. Several studies have demonstrated that PBDE metabolites compete with T4 for binding thyroid hormonetransport proteins (TTR, thyroxine-binding globulin [TBG]). Meerts et al. (1998, 2000) tested 17 PBDE congeners and 3 hydroxylated PBDEs for possible interaction with T4 binding to human TTR, a plasmatransport protein of thyroid hormones, in an in vitro competitive binding assay. None of the purecongeners competed with T4 for binding to human TTR without metabolic activation. Incubation of thecongeners with rat liver microsomes induced by PB (CYP2B enriched), β-napthoflavone (CYPIAenriched), or clofibrate (CYP4A3 enriched) indicated that 9 of the 17 pure congeners generatedmetabolites (not identified) that were able to displace T4 from TTR (>60% competition): BDE 15, BDE 28, BDE 30, BDE 47, BDE 51, BDE 75, BDE 77, BDE 100, and BDE 119. Testing of the threeknown hydroxylated PBDEs, used for their structural resemblance with the thyroid hormones3,5-diiodothyronine (3,5-T2), 3,3',5-triiodothyronine (T3), and 3,3',5,5'-tetraiodothyronine (T4) showed that the T4-like (2,6-dibromo-4-[2,4,6-tribromophenoxy]phenol) and T3-like (2-bromo-4-[2,4,6-tri PBDEs 2823. HEALTH EFFECTSbromophenoxy]phenol) hydroxylated PBDEs were 1.42- and 1.22-fold more potent, respectively, than T4; the T2-like hydroxylated PBDE (4-[2,4,6-tribromophenoxy]phenol) showed low affinity for TTR(0.41-fold less potent than T4). Consistent with these findings, Hamers et al. (2006) reported that 6OH BDE-47, but none of the 19 PBDE congeners, competed with the natural ligand T4 for TTR binding(4-fold less potent than T4). Additionally, Ren and Guo (2012) reported that 5 of the 11 OH-PBDEstested bound to TTR with 1.2–2-fold greater potency than T4 and 1 of the 11 OH-PBDEs tested bound to TBG with 1.5-fold greater potency than T4. In another study, all OH-PBDEs tested were consideredstrong binders of TTR (relative potency compared with T4 between 0.1 and 1) and moderate-to-strongbinders of TBR (relative potency compared with T4 between 0.01 and 1), including 3-OH-BDE 47, 5-OH-BDE 47, 6-OH-BDE 47, 4-OH-BDE 49, and 6-OH-BDE 99 (Marchesini et al. 2008). Parentcompounds (BDE 47, BDE 49, BDE 68, and BDE 99) and MeO-PBDEs (6-MeO BDE 57, 2’-MeO-BDE 68) showed no-to-weak binding (<0.01 relative potency compared with T4). Marsh et al.(1998) tested the affinity of 4'-hydroxyl derivatives of 1,3,5-triBDE, 1,3,3',5'-tetraBDE, and 1,3,3',5,5'-pentaBDE were tested for human thyroid hormone receptor subunits THR-α and THR-β in vitro. These congeners were tested because they theoretically show the highest structural similarity to T4 and T3. None of the hydroxylated derivatives effectively competed with the thyroid hormones for bindingto either receptor (affinities were 41–>1,000 times less than for T4 and T3).Studies also suggest that PBDEs and/or their metabolites can alter thyroid hormone binding to thyroidreceptors. In a reporter gene assay in Chinese hamster ovary (CHO) cells, 4OH-BDE-90 was antagonistic to both THR-α and THR-β receptors, with a potency 2-fold less potent in the THR-α assay and ~30%more potent in the THR-β assay than the reference compound tetrabrominated bisphenol A (TBBPA)(Kojima et al. 2009). However, receptor antagonism was not observed for the other three OH-PBDEstested (4OH-BDE-17, 4OH-BDE-42, 4OH-BDE-49) or any of the eight PBDEs or four MeO-PBDEstested (Kojima et al. 2009). In another reporter gene assay in fibroblast-derived CV-1 cells, T3-dependentTHR-α- and THR-β-responsive gene expression was decreased by BDE 100, BDE 153, BDE 154, BDE 290, and DE-71 in fibroblast-derived CV-1 cells by 35–45% (Ibhazehiebo et al. 2011). Thyroid responsive element (TRE) dissociation from TRs was also significantly increased by BDE 100, BDE 154, and BDE 290 by 30–45%; however, no changes in THR-α and THR-β cofactor recruitment were observed for 11 PBDEs, 2 OH-PBDEs, or DE-71 (Ibhazehiebo et al. 2011). In contrast, in a THR-α and THR-β cofactor recruitment assay, 4 of the 10 OH-PBDEs tested were TR agonists (2OH-BDE-28, 3OH-BDE-28, 5OH-BDE-47, 6OH-BDE-47), showing 70–90% of the maximal response induced by T3 (Ren et al. 2013). PBDEs 2833. HEALTH EFFECTSIn general, lower-brominated PBDEs, and their hydroxylated (OH-PBDEs) and methoyxylated (MeO-PBDEs) metabolites, show estrogenic activities 5–7 orders of magnitude lower than the naturalligand 17β-estradiol (E2) in reporter gene assays, and higher-brominated PBDEs and their metabolitesshow anti-estrogenic activities that are 1–6 orders of magnitude lower than reference antiestrogenic compounds (Hamers et al. 2006; Kojima et al. 2009; Li et al. 2013b; Meerts et al. 2001; Mercado-Feliciano and Bigsby 2008b). A good correlation has been shown between estrogenic activity and ERbinding affinity of low-brominated PBDEs (Li et al. 2013b; Mercado-Feliciano and Bigsby 2008b). Onestudy reported no estrogenic activity in an ER-dependent gene transcriptional activation assay for“PBDEs” in human HeLa 9903 cells stably transfected with the human ERα receptor; however, the specific congener(s) were not identified (Kim et al. 2011b).For two OH-PBDEs that have bromine substitution patterns similar to the thyroid hormones T2 [3,5-diiodothyronine] and T3 [3,3’,5-triiodothyronine] (i.e., 4-(2,4,6-tribromophenoxy)phenol and 2-bromo-4-(2,4,6-tribromophenoxy) phenol, respectively), estrogenic activities were 2–10 times morepotent than E2 (Meerts et al. 2001). The T2-like hydoxylated PBDE 4-(2,4,6-tribromophenoxy)phenolalso showed estrogenic activity in ERα- and ERβ-like human embryonic kidney cells, with maximuminductions of 50–95% of the maximum induction by E2 (Meerts et al. 2001). Estrogenic effects were also demonstrated using cell proliferation assays in breast cancer cells exposed tothe commercial pentaBDE mixture DE-71 (Mercado-Feliciano and Bigsby 2008b). Basal cellproliferation was significantly increased by up to 10-fold with DE-71 exposure (compared with increasesup to 15-fold with exposure to the natural ligand E2). Co-exposure of cells to DE-71 and E2 significantlydecreased proliferation relative to E2-only exposure in a concentration-dependent manner, suggesting an antagonistic effect of DE-71 on E2-induced cell proliferation (Mercado-Feliciano and Bigsby 2008b). However, neither basal nor E2-induced cell proliferation were altered in breast cancer cells exposed toBDE 47, BDE 99, BDE 100, or BDE 209 (Kwiecińska et al. 2011). Caspase-9 activity (apoptotic marker)was significantly decreased by ~40–60% in tests with all congeners compared with decreases of ~25%with exposure to E2 (Kwiecińska et al. 2011).Estrogenic effects have also been reported in other in vitro and in vivo assays, including increased uterine mRNA expression of ERα and ERβ in adult female offspring of rats given subcutaneous injections ofBDE 99 at 1 mg/kg/day (but not 10 mg/kg/day) from GD 10 to 18 (Ceccatelli et al. 2006); increaseduterine mRNA and protein levels of calbindin-D9k, a biomarker for estrogenic compounds, in immaturefemale rats exposed to BDE 47 at 50–200 mg/kg/day from PND 16 to 18 (Dang et al. 2007); increased PBDEs 2843. HEALTH EFFECTSgene and protein expression of ERα and ERβ in porcine ovarian follicles by 5-OH-BDE-47 and 6-OH-BDE-47 (Karpeta et al. 2014); inhibition of E2 metabolism in rat liver microsomes by11 OH-PBDEs and 1 MeO-BDE (Lai and Cai et al. 2012); inhibition of E2 sulfation (E2SULT) inV79 cells by 6-OH-BDE-47, BDE 19, BDE 47, and BDE 49 (Hamers et al. 2006); and increased production of vitellogenin in trout hepatocytes exposed to BDE 47, BDE 99, or BDE 205 (Nakari and Pessala 2005). However, estrogenic effects were not observed with exposure to BDE 47, BDE 99, orBDE 205 in recombinant yeast assays with human estrogen receptor (hER) (Nakari and Pessala 2005). Ina human cohort study, total serum PBDE in adult females were positively associated with ERα and ERβmRNA expression levels in the blood; however, BDE 47 serum levels were negatively associated withERα and ERβ mRNA expression levels (Karmaus et al. 2011). Similarly, ERβ gene and proteinexpression were decreased in porcine ovarian follicles exposed go BDE 47; however, no change wasobserved in ERα gene or protein expression levels (Karpeta et al. 2014).Anti-androgenic and anti-prostagenic activity was observed in reporter gene assays in human osteoblastcells following exposure to 16 different PBDE congeners, 6OH-BDE-47, and 2 commercial PBDEmixtures (DE-71, Octa LM) (Hamers et al. 2006). The most potent antiandrogens, BDE 19 and BDE 100,were 21 and 13 times less potent than the reference antiandrogenic drug flutamide. BDE 19 was also themost potent antiprostagen, and was 3 orders of magnitude lower in potency than the reference compound RU-486 (Hamers et al. 2006). Neither antiandrogenic nor antiprogestagenic activity were observed forBDE 169, BDE 206, or BDE 209 (Hamers et al. 2006). Antiandrogenic activity was also observed in twoof five PBDEs tested (BDE 47, BDE 100) and the commercial pentaBDE mixture DE-71 tested in human breast cells (Stoker et al. 2005), and in five of eight PBDEs tested, three of four OH-PBDEs tested, and four of four MeO PBDEs (all low-brominated) tested in CHO cells (Kojima et al. 2009). In contrast, Christen et al. (2010) reported that BDE 100 and BDE 155 enhanced DHT-dependent activation ofandrogen receptor (AR)-responsive gene expression in human breast cells (no other congeners tested). However, findings were not concentration-dependent, with the greatest enhancement (~50%) at 10 nM(highest concentration used was 10 µM) (Christen et al. 2010). This suggests that very lowconcentrations of PBDEs may be androgenic, while higher concentrations appear to be antiandrogenic. No changes in AR gene or protein expression were observed in porcine ovarian follicular cells exposed toBDE 47, 5-OH-BDE-47 or 6-OH-BDE-47 (Karpeta et al. 2014). Anti-progestagenic activity was alsoobserved in an in vivo study, where adult female offspring of rats given subcutaneous injections of PBDE99 at 1 or 10 mg/kg/day from GD 10 to 18 had significantly decreased uterine mRNA expression levels ofprogesterone receptor (Ceccatelli et al. 2006). PBDEs 2853. HEALTH EFFECTSThe ability of the commercial pentaBDE mixture DE-71 and the pure congeners, BDE 47, BDE 99, BDE 100, BDE 153, and BDE 154 to compete with 1.0 nM [3H]R1881 for binding to the rat AR wasevaluated using a cytosolic extract prepared from rat ventral prostate tissue (Stoker et al. 2005). Relative binding curves suggested that all of the compounds competed with [3H]R1881 for binding to the AR, with inhibition up to 80 and 98% for DE-71 and BDE 100, respectively. In order to determine if this inhibition was competitive, additional in vitro binding tests for BDE 100 were conducted to calculate the inhibition constant (Ki). These experiments showed that BDE 100 was a competitive inhibitor with a Ki value of1 µM (Stoker et al. 2005). Anti-glucocortogenic activity was observed in 3/8 PBDEs (PBDE 85, PBDE 99, PBDE 100), 2/4 OH-PBDEs (4OH-BDE-17, 4OH-BDE-49), and 1/4 MeO-PBDEs (4MeO-BDE-49) in aglucocorticoid receptor-mediated reporter gene assay with CHO cells (Kojima et al. 2009). Antiglucocortogenic potencies were 3 orders of magnitude lower than for the reference compound RU-486.In summary, the mechanistic studies show that PBDEs and/or their metabolites are capable of acting as thyroid hormone transporters or receptors and are weakly estrogenic, anti-androgenic, anti-prostagenic,and anti-glucocortogenic. However, these findings were not always consistent between differentcongeners, metabolites, and studies. Therefore, mechanisms of endocrine disruption by PBDEs have notbeen fully elucidated.Mechanisms of Neurotoxicity. As detailed in Section 3.2.2.4 Neurological Effects, developmentalexposure to PBDEs has been associated with altered neurodevelopment and behavior later in life in bothhumans and animals. The mechanisms for these behavioral and cognitive effects have not beenelucidated; however, proposed mechanisms include neuroendocrine disruption (including altered thyroid hormone homeostasis), alterations in neurotransmitter systems (cholinergic, dopaminergic, glutamatergic,and/or gabaergic), altered calcium homeostatic mechanisms, altered intracellular communication,oxidative stress, and cell death. Additionally, monohydroxylated metabolites are more potent than theparent BDE in several of the mechanistic assays, suggesting that bioactivation by oxidative metabolismcontributes to the neurotoxic potential of PBDEs.Since altered thyroid hormone levels have been reported in both animals and humans (see Section 3.2.2.2 Systemic Effects, Endocrine subsection), one possible mechanism of neurotoxicity involves the well-documented key role of thyroid hormones in brain development. In support of this mechanism, BDE 99 down-regulates the transcription of the thyroid receptors α1 and α2 (TRα, TRα2) in cultured rat cerebellar PBDEs 2863. HEALTH EFFECTSgranular cells, leading to disruption in the expression of T3-mediated genes, including decreased brain-derived neurotrophic factor (BDNF) (Blanco et al. 2011). BDE 99 and hydroxylated PBDE metabolites(3-OH-BDE-47, 6-OH-BDE-47, and 5’-OH-BDE-99) have also been shown to significantly decrease the activity of the selenodeiodinase type 2 iodothyronine deiodinase (DIO2), which converts T4 to T3 in thebrain, in primary human astrocytes and H4 glioma cells; BDE 47 did not affect DIO2 activity (Roberts etal. 2015). Decreased DIO2 activity associated with exposure to BDE 99 and OH-PBDEs was attributedto multiple mechanisms, including observed downregulation of DIO2 mRNA, competitive inhibition ofDIO2, and enhance post-translational degradation of DIO2. Ibhazehiebo et al. (2011) showed that theTR-antagonist, BDE 209 (see Mechanisms of Endocrine Disruption section above), inhibits T4-induced dendritic arborization in cultured rat cerebellar Purkinje cells. However, BDE 47, which was not found tobe a TR-antagonist (see Mechanisms of Endocrine Disruption section above), did not alter dendriticarborization (Ibhazehiebo et al. 2011). Study authors indicate that other PBDEs and OH-PBDEs were evaluated for dendritic developmental effects; however, results were not reported (Ibhazehiebo et al.2011). Additionally, Schreiber et al. (2010) showed that reduced migration and differentiation observed in cultured fetal human neural progenitor cells (hNPCs) following exposure to BDE 47 and BDE 99 wasprevented with co-exposure to T3. These findings indicate that in vitro neurodevelopmental changes were due to disruption of cellular thyroid hormone signaling. However, BDE 209 did not significantlyinterfere with the T3-mediated response in either a human embryonic kidney reporter cell line expressingmouse TRα (HEK293-Gal4TRα) or a mouse cerebellar neural reporter cell line expressing TRα(C17.2α-HRLuc), with or without metabolic activation (Guyot et al. 2014). Preliminary studies also did not show altered T3-mediated responses in C17.2α-HRLuc cells incubated with a commercial PBDEmixture (BDE-CM, AccuStandard, Inc.; according to manufacturer datasheet, the mixture contains equalparts BDE 28, 47, 99, 100, 153, 154, 183, and 209) (Guyot et al. 2014). Additionally, only slight, but statistically significant, changes were observed in the expression of T3-responsive genes in C17.2α cells following exposure to BDE 209 and T3, compared with T3 exposure alone (Guyot et al. 2014).Some studies suggest that the neurological effects of PBDEs might be related to alterations in cholinergicfunctions. For example, neonatal exposure to a single oral dose of BDE 99 (8 mg/kg) on PND 10 or asingle oral dose of BDE 209 (≥5.76 mg/kg) on PND 3 altered the behavioral response to cholinergic agents (nicotine or paraoxen) in adult mice (Buratovic et al. 2014; Johansson et al. 2008; Viberg et al.2002, 2007). Neonatal exposure to nicotine and adult exposure to BDE 99 (single 8 mg/kg oral dose atage 5 months) also affected behavior in mice, although the change was not seen in mice only exposed toBDE 99 as adults or mice only exposed to nicotine as neonates (Ankarberg et al. 2001). Additionally, thedensities of cholinergic nicotinic receptors in the brain hippocampus and/or cortex were decreased by 7– PBDEs 2873. HEALTH EFFECTS31% in adult mice exposed once to BDE 99 or BDE 153 at ≥0.8 mg/kg on PND 10 (Fischer et al. 2008; Viberg and Eriksson 2011;Viberg et al. 2003a, 2004b, 2005) and increased transcription of cholinergic receptors nAchR-β2, AChR5, and nAChR-α4 were observed in the cortex of 2-month-old mice exposed to BDE 99 at 12 mg/kg on PND 10 (Hallgren et al. 2015). However, no exposure-related effects were observed in any cholinergic parameters in 6- or 27-week-old mink from mink sows exposed to dietarypentaBDE at 0, 0.01, 0.05, or 0.25 mg/kg/day from pre-mating day 28 to PNW 6 (Bull et al. 2007). In ratneuroendocrine pheochromocytoma (PC12) cells and rat neuroblastoma (B35) cells, BDE 209 significantly decreased the Ach-evoked response in vitro (Hendriks et al. 2014). Additionally, in vitroexposure to BDE 99 led to altered neurotransmitter phenotype differentiation in rat PC12 cells, resultingin a decreased number of cholinergic cells and a greater number of dopaminergic cells; this was notobserved with exposure to BDE 47 (Dishaw et al. 2011; Slotkin et al. 2013). BDE 47 also did notmodulate human α4β2 nicotinic acetylcholine (nACh) receptor function (expressed in Xenopus oocytes); however, its hydroxylated metabolite, 6-OH-PBDE-47, was a nACh antagonist (Hendriks et al. 2010). Other studies have reported alterations in the dopaminergic system. Alterations have been observed in thestriatum of mice exposure to 30 mg/kg/day of the pentaBDE mixture DE-71 for 30 days via gavage,including reductions in dopamine levels, altered dopamine handling (i.e., altered dopamine breakdown into DOPAC and HVA), and reductions in the striatal dopamine transporter (DAT) and vesicularmonoamine transporter 2 (VMAT2) (Bradner et al. 2013). DE-71 also inhibited the in vitrouptake of the neurotransmitter dopamine into rat brain synaptic vesicles; however, inhibition was notobserved with commercial mixtures of octaBDE (DE-79) or decaBDE (DE-83R) (Mariussen and Fonnum2002, 2003; Mariussen et al. 2003). Similarly, DE-71 reduced synaptosomal dopamine concentrationsand increased medium dopamine concentrations in striatal synaptosomes derived from PND 7–21 rats(Dreiem et al. 2010). However, in contrast to the cholinergic system, acute neonatal exposure to BDE 99(12 mg/kg) on PND 10 did not significantly affect dopaminergic gene transcription in the cortex orhippocampus on PND 11 or at 2 months of age; genes evaluated included tyrosine hydroxylase (TH), DAT, and dopamine receptors D1, D2, and D5 (DRD1, DRD2, and DRD5) (Hallgren and Viberg 2016).Additional evidence suggests that alterations in the glutamatergic system may also contribute to observed neurological effects following PBDE exposure. Neonatal exposure to BDE 209 via gavage doses of20 mg/kg/day from PND 3 to 10 resulted in significant upregulation of NR1 mRNA in the frontal cortex and hippocampus of PND 11 and PND 60 mice, as well as decreased binding of the regulatory complexREST/NRST (Repressor Element Silencing Transcription Factor/Neuron-Restrictive Silencer Factor) tothe NR1 promotor (Verma et al. 2015). Additionally, extracellular glutamate levels (along with oxidative PBDEs 2883. HEALTH EFFECTSstress and cell death) were significantly elevated in cultured mouse cerebellar neurons following in vitro exposure to BDE 47 (Costa et al. 2016). Antagonists of ionotropic glutamate receptors, but notmetabotropic, prevented BDE 47-induced oxidative stress and cell death, providing additional evidencefor a role of glutamatergic signaling in PBDE neurotoxicity (Costa et al. 2016). In rats, exposure to BDE47 at doses ≥0.1 mg/kg/day for 30 days via gavage resulted in altered mRNA expression of the glutamate receptor subunits NR(1), NR(2)B, and NR(2)C (Yan et al. 2012).Limited evidence suggests that alterations in the gabaergic system may also contribute to observedneurological effects following PBDE exposure. A commercial pentaBDE mixture (DE-71) caused a slight inhibition of in vitro uptake of the neurotransmitter GABA into rat brain synaptic vesicles, butcommercial mixtures of octaBDE (DE-79) or decaBDE (DE-83R) did not alter GABA uptake (Mariussen and Fonnum 2003). Additionally, the metabolite, 6-OH-PBDE-47, was a partial agonist for the human GABAA receptor expressed in Xenopus oocytes; however, its parent compound (BDE 47) did notmodulate GABAA receptor activity (Hendriks et al. 2010).Effects of PBDEs on the function and development of the nervous system could also involve disruption ofcalcium homeostatic mechanisms and intracellular signaling events. In human neural progenitor cells(hNPCs), in vitro exposure to BDE 47 or 6-OH-BDE-47 led to transient increases in intracellular Ca+2 levels due to increased influx of extracellular Ca+2 as well as intracellular release from the endoplasmicreticulum (Gassmann et al. 2014). Based on additional studies using multiple inhibitors/stimulators ofpresumably involved signaling pathways, the increase in extracellular influx appears to be due to interference with the cell membrane, rather than alterations of Ca+ ion channel dynamics, and theincreased ER release was associated with activation of protein lipase C and inositol 1,4,5-trisphophatereceptor, independent of the ryanodine receptors (Gassmann et al. 2014). A series of studies evaluatedthe effects of in vitro exposure to PBDEs and their metabolites on calcium homeostasis in rat PC12 cells (Dingemans et al. 2008, 2010a, 2010b). Collectively, these studies show that hydroxylated metabolites ofPBDEs lead to increased basal Ca+2 levels due to Ca+2 release from the endoplasmic reticulum andmitochondria as well as decreased depolarization-evoked Ca+2 levels. BDE 47 was shown to havesimilar, but less potent effects, while no effects on calcium homeostasis were observed with BDE 49,BDE 99, BDE 100, BDE 153, or methylated metabolites (Dingemans et al. 2008, 2010a, 2010b). In asimilar study, BDE 209 did not alter calcium homeostasis in rat PC12 or B35 cells (Hendriks et al. 2014). BDE 49, OH-BDE-47, and 4’OH-BDE-49 have also been shown to be potent modulators of ryanodinereceptors type 1 and 2, which regulate essential aspects of Ca+2 signaling; BDE 47 was without activity inthis assay (Kim et al. 2011c; Pessah et al. 2010). In culture systems, BDE 47 and BDE 209 were shown PBDEs 2893. HEALTH EFFECTSto increased intracellular Ca+2 content in mouse cerebellar granule cells and rat neonatal hippocampalneurons, respectively (Chen et al. 2010; Costa et al. 2016). In vitro exposure to DE-71 or BDE 47 alsostimulated arachidonic acid release in rat cerebellar granule neurons; this effect was not seen with DE-79 (Kodavanti 2003; Kodavanti and Derr-Yellin 2002). The release of arachidonic acid appeared to be mediated by the activation of both Ca+2-dependent and Ca+2-independent cytosolic phospholipase A2. In vitro exposure to DE-71and BDE 47 also caused translocation of protein kinase C, as indicated byincreased phorbol ester binding; DE-79 did not induce this effect (Kodavanti and Derr-Yellin 2002; Rao et al. 2003). Other effects of penta mixture DE-71 and tetra congener BDE 47 included decreases in intracellular calcium buffering by microsomes and mitochondria (Kodavanti and Derr-Yellin 2002). Thetetra- congener BDE 47 was generally more potent than the DE-71 mixture (mainly comprised of tetra-and penta- congeners) in these tests. All commercial mixtures and congeners tested (DE-71, DE-79, BDE47, BDE 77, BDE 99, BDE 153) elevated phosphorylated extracellular signal-regulated kinase (pERK) ½, with congeners having a greater effect than mixtures (Fan et al. 2010). pERK ½ is a widely studiedMAPK cascade known to be involved in learning and memory. Neurotoxicity may be mediated by cell death, perhaps in response to oxidative stress, as increasedapoptosis and upregulation of apoptotic proteins and markers of oxidative have been observed in thehippocampus, cortex, and cerebellum of rats following developmental exposure to PBDEs (He et al.2009; Chen et al. 2014; Cheng et al. 2009; Costa et al. 2015; see Section 3.2.2.6, Developmental Effectsfor more details). In support, ROS were increased in rat cerebellar granular cells exposed to 25 µM BDE 99, and this increase was correlated with a decrease in the gene expression of the anti-apoptoticprotein Bcl-2 (Blanco et al. 2011). Another study found that DE-71 was more toxic than octa- and decacongeners in inducing cell death and free radical formation in cerebellar granule cells (Reistad et al.2002). In cultured rat cortical cells, a high concentration of BDE 99 (30 µM) induced cell death withoutany apparent increase in caspase-3 activity (Alm et al. 2008). BDE 47 also induced apoptosis in primarycultured rat hippocampal neurons; changes in oxidative stress parameters included increased ROS levels,malondialdehyde content, glutathione peroxidase levels and decreased glutathione and superoxidedismutase levels (He et al. 2008c). BDE 47-induced apoptosis in human SH-SY5Y was shown to bemediated via the mitochondrial p53 pathway, as evidenced by up-regulation of p53 and Bax, down-regulation of Bcl-2 and Bcl-2/Bax ratio, enhancement of Cyt c release from mitochondria into the cytosol, and activation of caspase-3, as well as by ultrastructural abnormalities of mitochondria (Zhang et al. 2013c). In mouse cerebral granule cells, PBDE exposure caused decreased cell viability, induced apoptotic cell death, and increased ROS and lipid peroxidation following exposure to BDE 47, BDE 99, BDE 100, BDE 153, and BDE 209 (Costas et al. 2015, 2016; Huang et al. 2010). Huang et al. (2010) PBDEs 2903. HEALTH EFFECTSreported a congener potency ranking of BDE 100 > BDE 47 > BDE 99 > BDE 153 >> BDE 209. Similarly, in cultured rat neonatal hippocampal neurons, BDE 209 decreased cell viability and increasedthe rate of apoptosis, ROS levels, malondialdehyde (MDA) content, and NO content (MAPKs) (Chen etal. 2010). In rat B35 cells, increased ROS generation in the absence of decreased cell viability was observed following exposure to BDE 209; no changes in ROS generation or cell viability were observed in rat PC12 cells (Hendriks et al. 2014).Proteomic and genomic studies have identified multiple systems/pathways in the brain that can be alteredfollowing PBDE exposure; however, strong conclusions regarding mechanisms of neurotoxicity fromthese studies cannot be made at this point. In laboratory animals, in vivo exposure to ≥12 mg/kg/day ofDE-71, BDE 99, BDE 203, BDE 206, or BDE 209 for 1–30 days has been shown to alter expression ofproteins involved in mediating GABA and glutamate neurotransmission in the frontal cortex; neuronalsurvival, growth, and synaptogenesis in the cortex, hippocampus, cerebellum, and striatum; cytoskeletalproteins in the cortex and hippocampus; oxidative stress and apoptosis in the cerebellum and hippocampus; and metabolism and energy production in the hippocampus and cerebellum (Alm et al.2008, 2006; Buratovic et al. 2014; Bradner et al. 2013; Kodavanti et al. 2015; Verma et al. 2015; Viberg 2009a, 2009b; Viberg and Eriksson 2011; Viberg et al. 2008). In mice exposed to 0.45 mg/kg/day for 28 days, combined analysis of proteomic and genomic data using biological network analysis indicateddisturbances in the following functional categories: nervous system development and function, neurological disease, and behavior (Rasinger et al. 2014). Gene ontogeny analysis showed significantchanges in mitochondrion morphogenesis, Wnt receptor signaling pathway, ʟ-glutamate transport, and calcium ion transport into cytosol, while proteomics identified differential expression of dynamin 1, calcium/calmodulin-dependent protein kinase II delta, and alpha 4a tubulin (Rasinger et al. 2014). Inculture systems, proteomic analysis of neonatal hippocampal neural stem/progenitor cells followingexposure to BDE 47 and/or BDE 209 showed differential expression of 19 proteins, including proteinsinvolved in metabolism, signal transduction, transcription, transport, and cell structures (Song et al. 2014). Key proteins showing downregulation were cofilin-1, which is associated with cell cycle and neuronal migration, and vimentin, which is implicated in nervous system repair mechanisms.Epigenetic changes may also contribute to neurotoxicity following exposure to PBDEs. Byun et al. (2015) examined this hypothesis in mice exposed perinatally to BDE 47 by evaluating DNA methylationpatterns in mitochondrial genes involved in respiration (cytochrome c oxidase I, II, and III), nucleargenome methylation markers (5hmc, L1Rn), and nuclear candidate genes related to behavioral and brainfunctions (BDNF, corticotropin releasing hormone receptor 1, melanocortin 2 receptor, nuclear receptor PBDEs 2913. HEALTH EFFECTSsubfamily 3 group C member 1, and alpha-synuclein). Significant changes in methylation were notobserve in DNA methylation patterns for the majority of genes examined, although slight (<2-fold), butstatistically significant, methylation decreases were observed in cytochrome c oxidase II, L1Rn, BDNF, and nuclear receptor subfamily 3 group C member 1 (NR3CL) in mice exposed to BDE 47, compared with control (Byun et al. 2015). However, global DNA methylation was significantly decreased infemale, but not male, offspring of mouse dams exposed to tetraBDE from premating day 28 to PND 21 (Woods et al. 2012). Mechanisms of Reproductive Toxicity. As discussed above (see Mechanisms of Endocrine Disruption), PBDEs have been shown to be weakly estrogenic and anti-androgenic. Additionally, alteredsteroidogenesis has been observed in vitro. Limited data are available regarding other potentialmechanisms of reproductive toxicity, such as altered placental function due to increased prostaglandinsynthesis and impaired mitochondrial function in sperm cells.Testosterone secretion was increased by up to 3-fold in cultured rat Leydig cells exposed to a mixture oftetra- and pentaBDEs (PBDE-710) (Wang et al. 2011c) or BDE 47 alone (Zhao et al. 2011). Thesestudies suggest that PBDEs may stimulate testosterone secretion by acting directly on Leydig cells toactivate the cAMP pathway and increased expression of steroidogenic acute regulatory protein (StAR), as evidence by increased gene and protein expression of StAR (Wang et al. 2011c, Zhao et al. 2011) as wellas increased intracellular levels of cAMP, increased PKAα nuclear translocation, and increased activity ofCYP 11A1 (Wang et al. 2011c). In support, the observed increase in testosterone secretion was blockedin the presence of the adenylyl cyclase inhibitor SQ22536 (Wang et al. 2011c). However, gene expression of StAR was not upregulated in mouse Leydig cells exposed to BDE 47 (testosterone secretionwas not evaluated) (Schang et al. 2016).Testosterone production was also significantly increased by ~2–4.5-fold in porcine ovarian folliclesexposed in vitro to BDE 47, BDE 99, or BDE 100 (Karpeta et al. 2011). Androstenedione (A4) and progesterone production was also significantly increased, but responses were not concentration-dependent. No exposure-related changes were observed in E2 production. Potential mechanisms underlying increased testosterone production were judged to be congener-specific. For BDE 47, observed increases in 17β-HSD protein expression and activity and decreases in CYP19 (aromatase) activitysuggest that increased testosterone production is due to increased conversion of A4 to testosterone by17β-HSD and decreased conversion of testosterone to E2 by CYP 19. For BDE 100, observed increasesin CYP17 protein expression and activity and decreases in CYP19 activity suggest increased testosterone PBDEs 2923. HEALTH EFFECTSproduction is due to increased conversion of progesterone to A4 by 17β-HSD (which is later convertedinto testosterone) and decreased conversion of testosterone to E2 by CYP 19. BDE 99 did not alterprotein expression or activity of 17β-HSD, CYP 17, or CYP 19. The study authors suggest that themechanism of action for BDE 99 is prior to the secretion of progesterone (i.e., StAR or 3β-HSD). Incontrast, OH-PBDEs (5OH-BDE-47, 6OH-BDE-47) did not alter testosterone or A4 production in porcine ovarian follicles; however, E2 production was significantly increased for both metabolites by 2.4– 2.9-fold (Karpeta et al. 2013). Additionally, the OH-PBDEs increased protein expression and activity ofCYP 19, suggesting that the increase in E2 may be due to increased aromatase-mediated conversion oftestosterone to E2. Similarly, changes in E2 secretion in mid-phase porcine luteal cells (increased) andlate-phase porcine luteal cells (decreased) following in vitro exposure to BDE 47 was significantlycorrelated with CYP 19 activity, although neither 5-OH-BDE-47 nor 6-OH-BDE-47 altered E2 secretion(Gregoraszczuk et al. 2015). In an in vivo study, CYP 19 activity in the ovaries of female rats was not altered with exposure to BDE 209 doses up to 60 mg/kg/day for 28 days via gavage; however, CYP 17 activity in the adrenal glands was significantly decreased by up to 97% in females, but not males (Van derVen et al. 2008a). Progesterone production was significantly increased by 1.3–2.1-fold in mid-phase porcine luteal cells exposed to BDE 47 (≥250 ng/mL) or BDE 99 (≥50 ng/mL) for 24 hours, but not 48 hours (Gregorasczuket al. 2012). Exposure to BDE 100 did not alter progesterone levels at concentrations up to 500 ng/mL ateither time point. Following exposure for 24 hours, cells exposed to BDE 99 also showed a significantincrease in the activity and protein level of 3β-HSD (which converts pregnenolone into progesterone); no changes in protein levels or activity of CYP11A1 (which converts 25-hydroxycholesterol into progesterone) were observed for any congener. Increased gene expression of 3β-HSD was also observed in mouse Leydig cells exposed to BDE 47; however, no exposure-related changes were observed inprogesterone secretion (Schang et al. 2016). Another study did not observe increased progesteroneproduction in early-, mid-, or late-phase porcine luteal cells exposed to BDE 47 for 24 hours at doses up to 50 ng/mL (lower than the lowest effective dose in the previous study) (Gregorasczuk et al. 2015). However, metabolites of BDE 47 (5-OH-BDE-47 and 6-OH-BDE-47) significantly decreasedprogesterone secretion in mid- and late-phase cells, respectively; 6-OH-BDE-47 also significantly inhibited 3β-HSD activity in late-phase cells (Gregorasczuk et al. 2015). Gregorasczuk et al. (2012) also measured activity of caspases 3, 8, and 9 in porcine luteal cells collected during the middle luteal phase following exposure to BDE 47, 99, or 100. Increased activation of both intrinsic (caspase 9) and extrinsic(caspase 8) apoptotic pathways was observed at 24 hours, with additional activation of caspase 3 at48 hours, was observed with all congeners, indicating premature apoptosis of middle luteal cells PBDEs 2933. HEALTH EFFECTS(apoptosis should occur during late luteal phase). The study authors proposed that the combined effects of altered steroidogenesis and premature apoptosis observed following in vitro exposure to PBDEs could negatively impact the estrous cycle by shortening the luteal phase, and potentially fecundability. Several studies have also evaluated steroidogenesis in human adrenocortical carcinoma cells followingexposure to PBDEs or their metabolites. The parent compound, BDE 47, did not alter testosterone, androstenedione, E2, pregnenolone, 17α-OH-pregnenolone, progesterone, or 17α-OH- progesterone levels in human adrenocortical carcinoma cells (van den Dungen et al. 2015). As observed in porcinefollicular cells, increased E2 production accompanied by increased CYP 19 gene expression and activitywas observed in human adrenocortical carcinoma cells following exposure to the methyoxylated metabolites 6Cl-2MeO-BDE-68 and 6MeO-BDE852/10 (He et al. 2008c). However, other tested metabolites (6Cl-2OH-BDE-7, 5Cl-6OH-BDE-47, 2MeO-BDE-28) showed decreased E2 production (Heet al. 2008c). Testosterone production was increased by 2/10 tested MeO-PBDEs (6MeO-BDE85, 6MeO-BDE137), but 0/10 tested OH-PBDEs (He et al. 2008c). The activity of the steroidogenic enzymeCYP 19 was also evaluated in cultured H295R human adrenocortical carcinoma cells exposed to19 PBDEs, 5 OH-PBDEs, and one methoxylated (MeO-)PBDE (Canton et al. 2005). Two lowbrominated PBDEs (BDE 19, BDE 28) induced CYP 19 by 200%, while two higher-brominated PBDEs(BDE 206, BDE 209) inhibited CYP 19 by 61–64%. 6OH-BDE-99 and 6MeO-BDE-47 also inhibited CYP 19 by 46–67%. The 6OH-BDE-47 metabolite showed minimal inhibition accompanied bycytotoxicity. In another study evaluating 11 OH-PBDEs and 11 MeO-PBDEs, all OH-PBDEs inhibited CYP 19 (Canton et al. 2008). The most potent aromatase inhibitors were 6OH-BDE-49 and 3OH-BDE-47, which decreased CYP 19 activity by 54 and 27%, respectively. No CYP 19 inhibition wasobserved with the MeO-PBDEs.Park and Loch-Caruso (2015) evaluated the effects of BDE 47 exposure on prostaglandin E2 synthesis in first trimester human extravillous trophoblast cells (HTR-8/SVneo), as prostaglandin E2 is a pro-inflammatory regulator of trophoblast cellular functions critical for successful placentation. Significantly increased prostaglandin E2 release was following in vitro exposure to BDE 47. mRNA expression ofenzymes involved in prostaglandin synthesis and catabolism were also significantly altered, includingincreases in prostaglandin synthase 2 and COX-2 (rate-limiting enzyme of prostaglandin synthesis) and decreased 15-hydroxyprostaglandin dehydrogenase and prostaglandin E synthase. Prostaglandin changeswere completely blocked following treatment with a COX-2 inhibitor, confirming that prostaglandin upregulation was COX-dependent. Additionally, significant ROS generation was observed followingBDE 47 exposure. The peroxyl radical scavenger, α-tocopherol, blocked both ROS generation and PBDEs 2943. HEALTH EFFECTSprostaglandin release; however, it had no effect on COX-2 mRNA levels. The study authors proposedthat generation of ROS in the placenta following BDE 47 exposure stimulates prostaglandin E2 production via post-translational modifications of a COX-2-dependant pathway, and may lead to placentaldysfunction (Park and Loch-Caruso 2015). Following 8-week oral exposure to BDE 47, increased apoptosis was observed in early leptotene spermatocytes of adult male rats (Huang et al. 2015). Proteomic analysis of testicular tissue indicateddifferential expression of 64 proteins, including 20 proteins related to apoptosis. Of these 15/20 apoptoticproteins were located in the mitochondria. Protein expression data from cultured testicular mouse cells (GC1-spg cells) were consistent with in vivo results, showing significant decreases in Uqcrc1, Atp5b, Tufm, Sucla2, and Lap3, indicating mitochondrial dysfunction and apoptosis, and a significant decreasein the anti-apoptotic factor Bcl-2 (Huang et al. 2015).Mechanisms of Immunotoxicity. As discussed in Section 3.2.2.3, Immunological and LymphoreticularEffects, limited evidence from animal studies suggest that PBDEs may cause immunosuppression(Darnerud and Thuvander 1998; Feng et al. 2016b; Fowles et al. 1994; Watanabe et al. 2008, 2010b). Lundgren et al. (2009) suggested that decreased immune response to infection may be due to suppression of cytokines by PBDE exposure. In mice infected with human coxsackievirus B3 (CVB3), serum levelsof IL-12, MIP-1β, RANTES, IFN-γ, and KC were markedly decreased (or completely lacking) in miceexposed to BDE 99 or Bromkal 70-5 DE (37% BDE 47, 35% BDE 99), compared with infected controls.No information regarding potential mechanisms of immunotoxicity for PBDEs were located. Howie et al.(1990) examined the immunotoxic potencies of various polychlorinated diphenyl ether congeners on theinhibition of the plaque-forming splenic cell response to sheep red blood cell antigen in mice. Theobserved potency order (2,3,3',4,4',5-hexaCDE > 3,3',4,4',5-pentaCDE > 2,3',4,4',5-pentaCDE >3,3',4,4'-tetraCDE > 2,2',4,4',5,5'-hexaCDE > 2,2',4,5,5'-pentaCDE > 2,2',4,4',5,6'-hexaCDE) generallyparalleled the congener-specific potencies for induction of hepatic microsomal AHH and EROD. Worth noting is the fact that the resulting ranking order of potency did not follow the order that would have been expected for a response known to be AhR-mediated, such as the inhibition of the plaque-forming spleniccell response to challenge with sheep red blood cells antigen. For example, the laterally substitutedcongeners 3,3',4,4'-tetraCDE and 3,3',4,4',5-pentaCDE were less immunotoxic than their respectivemonoortho-substituted analogs; this was true also for their enzyme induction potencies. These findingsshowed that increasing ortho- substitution is less effective in reducing the “dioxin-like” activity of thesecompounds. Howie et al. (1990) suggested that the ether bridge in the polychlorinated diphenyl ether PBDEs 2953. HEALTH EFFECTSmolecules increases the bond length between the two phenyl rings, thus diminishing the effects of orthosubstituents on the biochemical and toxic potencies of these compounds. However, these findings may ormay not be relevant to immunotoxic activities of PBDEs because of the considerable difference in themolecular size of brominated and chlorinated analogues, which may influence receptor-mediated effects,as well as potential toxicokinetic differences.Mechanisms of Hepatotoxicity. As discussed in Section 3.2.2.2 Systemic Effects (hepatic subsection),liver effects have been reported in adult, pregnant, and developing animals exposed to lower-brominated PBDEs. Evidence for hepatic effects in animals following exposure to decaBDE is less consistent.Existing information suggests that PBDEs can be classified as mixed-type inducers of hepatic microsomalmonooxygenases (see General Mechanisms of Toxicity Section). Observed effects of liver enlargementand hepatocellular hypertrophy are consistent with enzymatic induction. However, oxidative stress,inflammatory processes, and induction of apoptosis may also contribute to hepatotoxicity of PBDEs. In12-week gavage studies in mice, administration of troxerutin, an antioxidant and anti-inflammatory agent, prevented tetraBDE-induced hepatotoxicity (increased liver weight, increased serum ALT, hepatocytehypertrophy and vacuolization, and inflammatory cell infiltration) (Zhang et al. 2015a, 2015b).Moreover, tetraBDE-mediated induction of protein and genetic markers of oxidative stress, inflammation, proteasomal subunits, ER-stress pathways, and apopotisis are prevented by troxerutin. Zhang et al. (2015b) proposed that upregulation of NF-kB via the tetraBDE-induction of histone 3 acetylation at lysine 9 (H3K9) is due to tetraBDE-mediated decreases in SirT1 (which mediates deacetylation), and thatthis upregulation of NF-kB leads to the observed increases in mRNA levels of inflammation-relatedgenes. Additionally, Zhang et al. (2015a) suggested that oxidative stress leads to apoptosis via proteasome dysfunction-mediated ER stress. Taken together, the therapeutic mechanism of actions fortroxerutin could include antioxidant actions and restoration of proteasomal function, as well as the activation of SirT1. These are supported by similar decreases in markers of oxidative stress followingvitamin E treatment as well as the blunting of troxerutin-mediate effects with concurrent EX527injections (a SirT1-specific inhibitor) or epoxomicin injections (a selective proteasome inhibitor) (Zhanget al. 2015a, 2015b). In vitro studies also report elevated ROS levels, depleted GSH levels, mitochondrial damage and dysfunction, and apoptosis in HepG2 human hepatocellular carcinoma cells exposed to BDE47 (Liu et al. 2015; Saquib et al. 2016; Yeh et al. 2015). As observed in vivo, known antioxidants(eicosapentaenoic acid [EPA] and docosahexaenoic acid [DHA]) protect cultured HepG2 human hepatocellular carcinoma cells from BDE-47-induced oxidative stress and mitochondrial dysfunction (Yeh et al. 2015). However, Pereira et al. (2014) suggest that apoptosis may be mediated via direct PBDEs 2963. HEALTH EFFECTSinteraction of PBDEs with mitochondria, rather than secondary to PBDE-mediated oxidative stress. Inisolated rat liver mitochondria, BDE 154 directly interacted with the mitochondrial membrane, permeabilizing the membrane and leading to ATP depletion. These events occurred in the absence ofROS accumulation in the mitochondria, suggesting that BDE 154 impairs mitochondrial bioenergeticsand permeabilizes the mitochondria, potentially leading to cell death, in the absence of oxidative stress (Pereira et al. 2014). BDE 47 and BDE 99 have also been shown to directly interact with the mitochondrial membrane in isolated rat liver mitochondria, resulting in altered mitochondrialbioenergetics and ATP depletion (Pazin et al. 2015).3.5.3 Animal-to-Human Extrapolations Residues of PBDEs in humans reflect multiple exposure pathways and congener-specific elimination and thus, in general, represent steady-state body burdens that do not match the congener profiles in theoriginal exposure sources. For example, profiles of PBDE congeners in human milk do not resemble thepattern of any of the previously used commercial mixtures, which were predominantly pentaBDE, octaBDE, and/or decaBDE, as illustrated by the finding that the major PBDE congener in milk fromSwedish mothers was 2,2',4,4'-tetraBDE (BDE 47), which comprised approximately 55% of the totalPBDEs (Darnerud et al. 1998). As discussed in Chapter 6, residue analyses indicate that tetra- to hexacongeners predominate in humans, aquatic mammals, birds, fish, and other biota, indicating that thebiological fate of PBDE congeners is qualitatively similar in various animal species. The wildlife residue data also indicate that different species have different tissue ratios of congeners, possibly reflective ofinterspecies differences in metabolic capabilities as well as potential differences in exposure. The likelihood of interspecies differences in the quantitative disposition of PBDEs is illustrated by theobservation that metabolism and urinary excretion of a single oral dose of BDE 47 was significantly slower in rats than in mice (Orn and Klasson-Wehler 1998; Staskal et al. 2006b).While alterations in thyroid hormone levels are consistently altered in in PBDE-exposed animals, human data are less consistent. Humans are possibly less sensitive than rats to effects of PBDEs on circulatinglevels of thyroid hormones. This difference is thought to derive from the rat thyroid having a smallerstore of iodinated thyroglobulin that is more easily depleted when the availability of iodide is limited, andfrom a more rapid clearance of T4 from the rat circulation; the latter resulting from rats not having a highaffinity binding protein for T4 in serum analogous to TBG in humans (Capen 1997). If the production ofT4 and T3 is impaired sufficiently to deplete the thyroid of stored iodinated thyroglobulin, the thyroid cannot produce or secrete amounts of T4 and T3 needed to support physiological demands, circulating PBDEs 2973. HEALTH EFFECTSlevels of T4 (free T4) and T3 decrease, and a state of thyroid hormone insufficiency ensues. TTR is the major thyroid hormone binding protein in rats, but not in man. In most mammals, including humans, TBG is the principal thyroid hormone binding protein; 74% of the total bound-T4 is bound to TBG, and TTR and albumin bind only 11 and 15%, respectively, of the total (Schussler 2000). In contrast to mostmammals, the rat utilizes TTR as the major T4 plasma binding protein; approximately 75% of T4 in rat serum is bound to TTR and only 25% to albumin. Both circulating T3 and T4 are highly protein bound with only a small fraction of their total present as free hormone, and this high degree of protein bindingserves to maintain equilibrium between the extracellular and intracellular pools of these hormones(O’Connor et al. 1999). Less is known about the relative sensitivities of humans and experimental animals to developmentaleffects of PBDEs. Outstanding uncertainties include potential differences in kinetics of maternal-fetal and maternal-infant transfer of PBDEs, as well as potential differences in the degree to which the fetus ofthe human, in comparison to experimental animals, is dependent on maternal thyroid hormone fordevelopment, particularly during the period of gestation prior to the onset of fetal hormone production.3.6 TOXICITIES MEDIATED THROUGH THE NEUROENDOCRINE AXISRecently, attention has focused on the potential hazardous effects of certain chemicals on the endocrinesystem because of the ability of these chemicals to mimic or block endogenous hormones. Chemicals with this type of activity are most commonly referred to as endocrine disruptors. However, appropriateterminology to describe such effects remains controversial. The terminology endocrine disruptors, initially used by Thomas and Colborn (1992), was also used in 1996 when Congress mandated the EPA to develop a screening program for “...certain substances [which] may have an effect produced by a naturally occurring estrogen, or other such endocrine effect[s]...”. To meet this mandate, EPA convened a panel called the Endocrine Disruptors Screening and Testing Advisory Committee (EDSTAC), and in 1998, the EDSTAC completed its deliberations and made recommendations to EPA concerning endocrinedisruptors. In 1999, the National Academy of Sciences released a report that referred to these same types of chemicals as hormonally active agents. The terminology endocrine modulators has also been used toconvey the fact that effects caused by such chemicals may not necessarily be adverse. Many scientists agree that chemicals with the ability to disrupt or modulate the endocrine system are a potential threat tothe health of humans, aquatic animals, and wildlife. However, others think that endocrine-active chemicals do not pose a significant health risk, particularly in view of the fact that hormone mimics existin the natural environment. Examples of natural hormone mimics are the isoflavinoid phytoestrogens PBDEs 2983. HEALTH EFFECTS(Adlercreutz 1995; Livingston 1978; Mayr et al. 1992). These chemicals are derived from plants and are similar in structure and action to endogenous estrogen. Although the public health significance and descriptive terminology of substances capable of affecting the endocrine system remains controversial,scientists agree that these chemicals may affect the synthesis, secretion, transport, binding, action, orelimination of natural hormones in the body responsible for maintaining homeostasis, reproduction, development, and/or behavior (EPA 1997). Stated differently, such compounds may cause toxicities thatare mediated through the neuroendocrine axis. As a result, these chemicals may play a role in altering, for example, metabolic, sexual, immune, and neurobehavioral function. Such chemicals are also thoughtto be involved in inducing breast, testicular, and prostate cancers, as well as endometriosis (Berger 1994;Giwercman et al. 1993; Hoel et al. 1992).Concern has been raised that many industrial chemicals, including PBDEs, are endocrine-active compounds capable of having widespread effects on humans and wildlife (Colborn et al. 1993; Crisp et al.1998; Daston et al. 1997; Safe and Zacharewski 1997). Particular attention has been paid to thepossibility of these compounds mimicking or antagonizing the action of estrogen. Estrogen influences thegrowth, differentiation, and functioning of many target tissues, including female and male reproductivesystems, such as mammary gland, uterus, vagina, ovary, testes, epididymis, and prostate. In addition, there is evidence that some of these environmentally-persistent chemicals alter the thyroid hormonesystem, which is a very important system for normal structural and functional development of sexualorgans and the brain. As discussed in Chapter 2 and Section 3.2.2.2. (Systemic Effects), the thyroid is a target of concernfollowing exposure to PBDEs, with adverse effects including altered thyroid hormone expression (predominantly reduced serum T4) and histological changes in the thyroid indicative of glandularstimulation (e.g., follicular cell hyperplasia similar to that induced by a hypothyroid state). Additionally, as discussed in Chapter 2 and Section 3.2.2.6 (Developmental Effects), the developing nervous systemand the developing reproductive system are also targets of concern. Since these effects could be mediatedby the neuroendocrine axis, several studies have tested PBDEs and their metabolites in in vitro endocrinedisruption screens and in vivo gene expression assays. These studies, and their results, are summarized in Section 3.5.2 (Mechanisms of Toxicity) in the subsections on Mechanisms of Endocrine Disruption (thyroid hormone assays; anti-estrogenic, -androgenic, -progestagenic, and -glucocortogenic assays) andMechanisms of Reproductive Toxicity (steroidogenesis assays). While results are not always consistentbetween studies, the data collectively indicate that there is a potential for some PBDEs to disrupt thyroidand other endocrine system functions in humans. PBDEs 2993. HEALTH EFFECTSAdditionally, a few studies specifically evaluated neuroendocrine disruption in vitro. BDE 47 and BDE 99 altered neurodevelopment of cultured fetal hNPCs (Schreiber et al. 2010). The migrationdistance of the hNPCs was reduced by up to 30% by BDE 47 and up to 35% by BDE 99. Differentiation of hNPCs into neurons was reduced up to 50% by BDE 47 and up to 68% by BDE 99, and differentiation of hNPCs into oligodendrocytes was reduced up to 51% by BDE 47 and up to 93% by BDE 99. Co-exposure with T3 prevented the effects of PBDE exposure on migration and differentiation, indicating thatneurodevelopmental changes were due to endocrine disruption of cellular thyroid hormone signaling. Similarly, Ibhazehiebo et al. (2011) showed that the TR-antagonist, BDE 209 (see discussion in Mechanisms of Endocrine Disruption), inhibits T4-induced dendritic arborization in cultured rat cerebellarPurkinje cells. However, BDE 47, which was not found to be a TR-antagonist (see discussion in Mechanisms of Endocrine Disruption), did not alter dendritic arborization (Ibhazehiebo et al. 2011). Study authors indicate that other PBDEs and OH-PBDEs were evaluated for dendritic developmentaleffects, but results were not reported (Ibhazehiebo et al. 2011). In another study, hyperosmotic-stimulatedsomatodendritic vasopressin release in supraoptic tissue from the hypothalamus of rats was significantlyreduced by 40–50% by BDE 46, BDE 77, and the commercial pentaBDE mixture DE-71 (no otherPBDEs tested) (Coburn et al. 2007). The neuropeptide vasopressin, which is synthesized in magnocellular neuroendocrine cells, functions to maintain body fluid homeostasis, cardiovascular control, learning and memory, and nervous system development.3.7 CHILDREN’S SUSCEPTIBILITYThis section discusses potential health effects from exposures during the period from conception to maturity at 18 years of age in humans, when all biological systems will have fully developed. Potentialeffects on offspring resulting from exposures of parental germ cells are considered, as well as any indirecteffects on the fetus and neonate resulting from maternal exposure during gestation and lactation. Relevant animal and in vitro models are also discussed.Children are not small adults. They differ from adults in their exposures and may differ in theirsusceptibility to hazardous chemicals. Children’s unique physiology and behavior can influence theextent of their exposure. Exposures of children are discussed in Section 6.6, Exposures of Children.Children sometimes differ from adults in their susceptibility to hazardous chemicals, but whether there is a difference depends on the chemical (Guzelian et al. 1992; NRC 1993). Children may be more or less PBDEs 3003. HEALTH EFFECTSsusceptible than adults to health effects, and the relationship may change with developmental age (Guzelian et al. 1992; NRC 1993). Vulnerability often depends on developmental stage. There arecritical periods of structural and functional development during both prenatal and postnatal life, and aparticular structure or function will be most sensitive to disruption during its critical period(s). Damagemay not be evident until a later stage of development. There are often differences in pharmacokinetics and metabolism between children and adults. For example, absorption may be different in neonatesbecause of the immaturity of their gastrointestinal tract and their larger skin surface area in proportion tobody weight (Morselli et al. 1980; NRC 1993); the gastrointestinal absorption of lead is greatest in infants and young children (Ziegler et al. 1978). Distribution of xenobiotics may be different; for example, infants have a larger proportion of their bodies as extracellular water, and their brains and livers are proportionately larger (Altman and Dittmer 1974; Fomon 1966; Fomon et al. 1982; Owen and Brozek1966; Widdowson and Dickerson 1964). The fetus/infant has an immature (developing) blood-brain barrier that past literature has often described as being leaky and poorly intact (Costa et al. 2004). However, current evidence suggests that the blood-brain barrier is anatomically and physically intact at this stage of development, and the restrictive intracellular junctions that exist at the blood-CNS interface are fully formed, intact, and functionally effective (Saunders et al. 2008, 2012).However, during development of the blood-brain barrier, there are differences between fetuses/infants andadults which are toxicologically important. These differences mainly involve variations in physiological transport systems that form during development (Ek et al. 2012). These transport mechanisms (influx and efflux) play an important role in the movement of amino acids and other vital substances across theblood-brain barrier in the developing brain; these transport mechanisms are far more active in the developing brain than in the adult. Because many drugs or potential toxins may be transported into thebrain using these same transport mechanisms—the developing brain may be rendered more vulnerable than the adult. Thus, concern regarding possible involvement of the blood-brain barrier with enhanced susceptibility of the developing brain to toxins is valid. It is important to note however, that this potentialselective vulnerability of the developing brain is associated with essential normal physiologicalmechanisms; and not because of an absence or deficiency of anatomical/physical barrier mechanisms.The presence of these unique transport systems in the developing brain of the fetus/infant is intriguing; asit raises a very important toxicological question as to whether these mechanisms provide protection forthe developing brain or do they render it more vulnerable to toxic injury. Each case of chemical exposureshould be assessed on a case-by-case basis. Research continues into the function and structure of the blood-brain barrier in early life (Kearns et al. 2003; Saunders et al. 2012; Scheuplein et al. 2002). PBDEs 3013. HEALTH EFFECTSMany xenobiotic metabolizing enzymes have distinctive developmental patterns. At various stages ofgrowth and development, levels of particular enzymes may be higher or lower than those of adults, and sometimes unique enzymes may exist at particular developmental stages (Komori et al. 1990; Leeder andKearns 1997; NRC 1993; Vieira et al. 1996). Whether differences in xenobiotic metabolism make thechild more or less susceptible also depends on whether the relevant enzymes are involved in activation ofthe parent compound to its toxic form or in detoxification. There may also be differences in excretion,particularly in newborns who all have a low glomerular filtration rate and have not developed efficienttubular secretion and resorption capacities (Altman and Dittmer 1974; NRC 1993; West et al. 1948). Children and adults may differ in their capacity to repair damage from chemical insults. Children alsohave a longer remaining lifetime in which to express damage from chemicals; this potential is particularlyrelevant to cancer.Certain characteristics of the developing human may increase exposure or susceptibility, whereas others may decrease susceptibility to the same chemical. For example, although infants breathe more air perkilogram of body weight than adults breathe, this difference might be somewhat counterbalanced by theiralveoli being less developed, which results in a disproportionately smaller surface area for alveolarabsorption (NRC 1993).Body burden data, as well as intake modeling, suggest that infants and toddlers have higher exposures toPBDEs as compared to older children or adults (EPA 2010; Lorber 2008; Trudel et al. 2011; Wong et al. 2013). Exposure during development may occur by transfer of PBDEs that have accumulated in women’sbodies to the fetus across the placenta (Antignac et al. 2009, 2008; Chen et al. 2013; Kawashiro et al. 2008; Li et al. 2013a; Mazdai et al. 2003; Meijer et al. 2008; Qiu et al. 2009; Vizcaino et al. 2011; Wan etal. 2010). Placental transfer, although it may be limited in absolute amounts (Frederiksene et al. 2010), isa concern because of possible effects of PBDEs on sensitive immature tissues, organs, and systems, withpotentially serious long-lasting consequences. Because PBDEs are lipophilic substances, they canadditionally accumulate in breast milk and be transferred to nursing infants (Antignac et al. 2008, 2009; Malarvannan et al. 2013; Park et al. 2011; Schecter et al. 2010). Transfer of PBDEs via breast milk could be considerable and, like prenatal exposure, has the potential to contribute to altered development. Toddlers and older children are exposed to PBDEs in the same manner as the general population,primarily via ingestion of contaminated dust and consumption of contaminated foods (EPA 2010; Lorber2008). However, exposure from these sources may be greater in young children due to: (1) greater hand-to-mouth behavior, increasing the risk of ingestion of contaminated dust and/or residues from PBDE PBDEs 3023. HEALTH EFFECTStreated materials; and (2) increased exposure relative to their body weight via diet due to increased caloricintake normalized to body weight and a higher general intake of animal fats. Additionally, children may be particularly vulnerable to PBDEs because, compared to adults, they are growing more rapidly and are generally expected to have lower and distinct profiles of biotransformation enzymes, as well as muchsmaller fat depots for sequestering these lipophilic chemicals.In human studies, significant associations between umbilical cord and placental PBDE (BDE 28, BDE 47, BDE 99, BDE 153, BDE 183, and total) and adverse birth outcomes were reported in a comparison of128 normal births and 25 cases of adverse birth outcomes (low birth weight, premature birth, still birth)(Wu et al. 2010; Xu et al. 2015a). Elevated maternal serum BDE 47 levels were also significantlyassociated with an increased risk of preterm birth in a case-control study of 197 full-term births and 82 pre-term births in Nashville, Tennessee (Peltier et al. 2015). A study of births from 20 healthypregnant women in Taiwan found that elevated PBDE concentrations (BDE 47, BDE 99, BDE 100, and BDE 209) in breast milk was associated with significantly reduced infant birth weight, length, chestcircumference, and Quetelet’s index (i.e., BMI) (Chao et al. 2007). A significant negative association between breast milk PBDE concentrations (BDE 47, BDE 99, BDE 100, BDE 153, and their sum) and birth weight was also identified in a Swedish cohort of 254 women with median total PBDE levels of2.4 ng/g lipid (Lignell et al. 2013). Several studies have also reported a negative association between one or more PBDE congeners in maternal serum and birth weight, length, and/or head circumference: a prospective reproduction study of 234 couples in Michigan and Texas (Robledo et al. 2015a), a birth cohort of 215 Chinese women with a median total PBDE level of 21.68 ng/g lipid (Chen et al. 2015), and a birth cohort of 686 Spanish women (Lopez-Espinosa et al. 2015). In other studies, no significantassociations were observed between birth weight, length, or head circumference and maternal or cordserum PBDE concentrations (Foster et al. 2011; Harley et al. 2011; Mazdai et al. 2003; Miranda et al.2015; Serme-Gbedo et al. 2016; Tan et al. 2009). In another study, a positive association between birth weight and length and colostrum PBDE (BDE 47, BDE 99, BDE 100, BDE 153, and their sum) wasobserved in a Tanzanian cohort of 95 women with a median total PBDE concentration of 19.8 ng/g lipid (Müller et al. 2016). In general, available data from animal studies do not indicate that PBDEs are embryotoxic or fetotoxic atPBDE doses below doses that elicited maternal toxicity, although occasional observations of reduced pup weight were reported (Argus Research Laboratories 1985a, 1985b; Biesemeier et al. 2011; Bowers et al.2015; Branchi et al. 2001, 2002, 2005; Breslin et al. 1989; Ellis-Hutchings et al. 2009; Fujimoto et al. 2011; Hardy et al. 2001, 2002; Kodavanti et al. 2010; Koenig et al. 2012; Life Science Research Israel PBDEs 3033. HEALTH EFFECTSLtd. 1987; Poon et al. 2011; Rice et al. 2007; Saegusa et al. 2012; Ta et al. 2011; Talsness et al. 2005, 2008; Tseng et al. 2008, 2013; Wang et al. 2011a; Watanabe et al. 2008, 2010b; WIL Research Laboratories 1986; Woods et al. 2012; Zhou et al. 2002). At PBDE doses that elicited maternal toxicity(15 mg/kg/day in rabbits, 50 mg/kg/day in rats), developmental effects observed included increased post-implantation loss due to late resorptions and skeletal variations commonly associated with maternal toxicity (Breslin et al. 1989; WIL Research Laboratories 1986). A few studies did report effects at dosesbelow maternal toxicity, including delayed ossification and an increased incidence of internal variationsin GD 20 rat fetuses exposed to pentaBDE at 2 mg/kg/day from GD 6 to 19 (Blanco et al. 2012), increased postimplantation loss and number of resorptions and decreased number of live fetuses/litter in mice exposed to decaBDE at ≥750 mg/kg/day from GD 7 to 9 (Chi et al. 2011), and subcutaneous edemaand delayed ossification in rats exposed to a low purity (77%) decaBDE mixture at 1,000 mg/kg/day fromGD 6 to 15 (Dow Chemical Co. 1985; Norris et al. 1975a).A major target of concern in infants and children is the developing nervous system. Numerous studies have reported results suggestive of an effect of PBDE on neurodevelopment in children. PBDE levels incord blood, maternal or infant serum, and/or breast milk have been correlated with cognitive score andadaptive behavior deficits in infants (Chao et al. 2011; Gascon et al. 2012; Shy et al. 2011); mental and physical development deficits in infants/toddlers at ages 12, 24, and 36 months (Herbstman et al. 2010);language and social developmental score deficits in toddlers at 24 months (Ding et al. 2015); increasedimpulsivity in toddlers at 24–36 months (Adgent et al. 2014; Hoffman et al. 2012); poor socialcompetence and ADHD or increased attention problems in 4-year-old children (Cowell et al. 2015;Gascon et al. 2011); decreased IQ and increased hyperactivity in 5-year-old children (Chen et al. 2014);impaired fine motor coordination, verbal memory and comprehension, and sustained attention in 5– 7-year-old children (Eskenazi et al. 2013; Roze et al. 2009); and poor attention and executive function deficits in 9–12-year-old children (Sagiv et al. 2015). In one birth cohort, no associations were observed between maternal serum PBDEs and neonatal behavior in 5-week-old infants (Donauer et al. 2015) orautistic behaviors at 4–5-year-old children (Braun et al. 2014); however, children from the same cohortshowed associations between maternal serum PBDEs and decreased IQ and increased hyperactivity at5 years of age (Chen et al. 2014) and executive function deficits at 5–8 years of age (Vuong et al. 2016a). Pre- and peri-natal studies in animals also consistently reported neurodevelopmental effects followingexposure to lower-brominated PBDEs and decaBDE at doses ≥0.06 and ≥2.22 mg/kg/day, respectively, including neurobehavioral alterations, delayed ontogeny of reflexes, ultrastructural changes, alterednicotinic receptor density, altered electrophysiology, and altered gene and protein expression levels(Biesemeier et al. 2011; Blanco et al. 2013; Bowers et al. 2015; Branchi et al. 2001, 2002, 2005; PBDEs 3043. HEALTH EFFECTSBuratovic et al. 2014; Chen et al. 2014; Cheng et al. 2009; Eriksson et al. 2001, 2002b, 2006; Fischer etal. 2008; Fujimoto et al. 2011; Gee and Moser 2008; He et al. 2009, 2011; Johansson et al. 2008; Koeniget al. 2012; Kuriyama et al. 2004, 2005; Rice et al. 2007; Sand et al. 2004; Ta et al. 2011; Viberg et al. 2002, 2003a, 2003b, 2004a, 2004b, 2005, 2006, 2007; Woods et al. 2012; Xing et al. 2009). Consistentreports of neurological effects have not been found in adult animal studies, and the only studies that did report neurological effects in adults observed altered neurobehavior at doses above those that cause neurodevelopmental effects in pre- and peri-natal studies (26.2 mg/kg/day of pentaBDE and ≥0.1 mg/kg/day of tetraBDE) (Driscoll et al. 2009; Yan et al. 2012).Neurodevelopmental effects may be mediated through the neuroendocrine axis, as thyroid hormonesregulate cell proliferation, migration, and differentiation during development, and maintenance of normal levels is essential to normal growth and development. In support, neuronal migration and differentiation of fetal human neural progenitor cells (hNPCs) were significantly impaired following in vitro exposure totetraBDE (Schreiber et al. 2010), and decaBDE was shown to be a thyroid hormone receptor antagonistthat significantly inhibited T4-induced dendritic arborization in cultured rat cerebellar Purkinje cells(Ibhazehiebo et al. 2011). Additionally, PBDE-induced alterations in thyroid hormone binding to transport proteins and receptors have been demonstrated in several in vitro studies (Hamers et al. 2006;Ibhazehiebo et al. 2011; Marsh et al. 1998; Meerts et al. 1998, 2000; Ren and Guo 2012; Ren et al. 2013). In humans, evidence for thyroid hormone disruption in infants is inconclusive. Some studies reportednegative associations between developmental exposure to PBDEs and infant serum or cord blood T4 levels (Abdelouahab et al. 2013; Herbstman et al. 2010; Kim et al. 2011a); however, other studies reported no association (Kim et al. 2011d, 2012a, 2012b, 2015; Lignell et al. 2016; Lin et al. 2011;Mazdai et al. 2003; Shy et al. 2012). Associations between developmental PBDE exposure and infantserum or cord blood T3 and TSH were similarly inconsistent (Abdelouahab et al. 2013; Eggesbo et al.2011; Kim et al. 2011d, 2012a, 2012b, 2015; Lignell et al. 2016; Lin et al. 2011; Mazdai et al. 2003; Shyet al. 2012; Stapleton et al. 2011; Turyk et al. 2008). In contrast, reduced serum T4 levels have beenconsistently reported in animals exposed to lower-brominated PBDEs during development (Bansal et al.2014; Blanco et al. 2013; Bondy et al. 2011, 2013; Bowers et al. 2015; Ellis-Hutchings et al. 2006;Kodavanti et al. 2010; Kuriyama et al. 2007; Miller et al. 2012; Poon et al. 2011; Shah et al. 2011;Skarman et al. 2005; Szabo et al. 2009; Wang et al. 2011a; Zhang et al. 2009; Zhou et al. 2002). Somestudies also reported reduced serum T3 levels, although findings are less consistent (Blanco et al. 2013;Bondy et al. 2013; Bowers et al. 2015; Shah et al. 2011; Zhang et al. 2009). Consistent changes inthyroid hormones were not found in animals exposed to decaBDE during development (Fujimoto et al. 2011; Rice et al. 2007; Tseng et al. 2008). PBDEs 3053. HEALTH EFFECTSDisruption of circulating hormone levels can have markedly different effects, depending on the stage ofdevelopment, and even transient disruptions can produce permanent effects, such as mental retardation, impaired motor skills, and hearing and speech impediments (Boyages 2000; Fisher and Brown 2000). Several factors might contribute to a high vulnerability of the fetus and neonate to PBDEs. Relativelybrief periods of thyroid hormone insufficiency (e.g., 14 days) can produce measurable neurological deficits in newborn infants (van Vliet 1999). Furthermore, unlike the adult thyroid gland, which containsa relatively large store of T4 that is sufficient to support circulating levels of hormone for several months, the neonatal thyroid contains only enough hormone to support circulating levels of hormone for ≥1 day (van den Hove et al. 1999; Vulsma et al. 1989). Thus, even acute exposures to a dose of lowerbrominated PBDEs sufficient to suppress thyroid hormone production could potentially result in thyroid insufficiency in the neonate. The absorbed dose of lower-brominated PBDEs per unit of body mass is also likely to be higher in infants compared to adults exposed to similar levels of PBDEs because ofhigher intakes per unit of body mass and exposure from breast milk. It should be noted that screening ofall newborn children for hypothyroidism is already a widely accepted and legislatively mandated practice (LaFranchi 1999; Landenson et al. 2000). Newborns are tested for thyroid hormone levels within the first few days of life in the United States and most other developed countries, and treatment is startedimmediately if indicated (LaFranchi 1999; Landenson et al. 2000).The human relevance of the thyroid effects of lower-brominated BDEs in animals is unclear. Humans are generally regarded as being less sensitive than rats to effects of PBDEs on circulating thyroid hormones.This is thought to be related to a smaller and more rapid turnover of the hormone pool in the rat thyroid, and to a more rapid clearance of secreted hormone in the rat; the latter being, in part, related to theabsence of TBG in rats (Capen 1997). TTR is the major thyroid hormone binding protein in rats, whereasTBG is the main binding protein in man and most other mammals. However, although TTR is a minorthyroid hormone binding protein in humans, it is the principal protein involved in T4 transport to the brain in both rats and man (Blay et al. 1993; Sinjari et al. 1998). TTR does not transport T4 from thebloodstream to the brain, but rather is the main T4 binding protein in cerebral spinal fluid (CSF) in ratsand humans. In the rat, T4 is transported to the brain primarily through the blood-brain barrier, and notvia the choroid plexus and CSF (Blay et al. 1993). Also, the mechanism by which lower-brominatedBDEs cause decreased serum T4 might involve hepatic microsomal enzyme induction and consequentincreased metabolic formation of hydroxy signaling-metabolites, but humans are not particularly sensitive to this effect. PBDEs 3063. HEALTH EFFECTSBased on limited findings from human and animal studies, another target of concern in infants and children exposed to PBDEs is the developing reproductive system. Main et al. (2007) found a significantpositive relationship between levels of PBDE in breast milk and congenital cryptorchidism (undescended testes) in male offspring. The study compared levels of 14 PBDE congeners in breast milk of mothers of62 Danish and Finnish boys with cryptorchidism to mothers of 68 controls from the same population. Significant elevation of concentrations in cases versus controls were seen for the sum of all 14 congeners, for the sum of the 7 most prevalent congeners found in all mothers (BDE 28, BDE 47, BDE 66, BDE 99, BDE 100, BDE 153, BDE 154), and for 5 of the 7 most prevalent congeners. No significant associationswere observed between PBDE concentrations in children’s adipose tissue and cryptorchidism(Koskenniemi et al. 2015). Other studies of reproductive development found no relationship between concentrations of individual PBDE congeners (BDE 28, BDE 47, BDE 99, BDE 100, BDE 153) in mid-pregnancy serum samples from California mothers and hypospadias in their male offspring (Carmichaelet al. 2010) and no relationship between current serum levels of PBDE and various measures of sexualmaturation (e.g., age at thelarche, current breast development, age at menarche) in a small cohort of teen-aged Dutch children (Leijs et al. 2008). However, serum PBDE levels in 6–8-year-old females were significantly associated with delayed onset of puberty in a longitudinal cohort of U.S. girls (Windham etal. 2015a). A study of 55 Dutch boys found significant positive associations between maternal serumlevels of BDE 154 collected on week 35 of pregnancy (but not other congeners measured, includingBDE 47 and BDE 153, that occurred at higher levels) and serum levels of the sex hormones, E2, free E2, and inhibin B (but not testosterone, LH, FSH, or sex hormone binding globulin) in the baby boys at3 months of age and testes volume in the boys at 18 months of age, but no effect on penile length at eitherage (Meijer et al. 2012). Additionally, Warembourg et al. (2016) reported an inverse association between cord serum BDE 209 and total testosterone (but not free testosterone, E2, aromatase index, sex hormonebinding globulin, or Anti-Müllerian hormone) in 141 French boys.In animal studies, reproductive effects were observed in adult F1 offspring exposed to a single dose ofpentaBDE at 0.06 mg/kg on GD 6, including reductions in testicular weight, sperm/spermatid number, and daily sperm production in males and a decreased number of secondary follicles and ultrastructuralchanges in the ovaries in females (although F1 fertility when mated to an unexposed animal was notimpaired) (Kuriyama et al. 2005; Talsness et al. 2005, 2008). In animals exposed pre- or perinatally todecaBDE, one study reported reproductive effects in adult male offspring exposed to decaBDE doses of10–1,500 mg/kg/day from GD 0 to 17, including testicular lesions, decreased AGD, and altered spermparameters (Tseng et al. 2013); however, no exposure-related changes in AGD, onset of puberty, orreproductive organ weight and histology were reported in offspring exposed to decaBDE at doses up to PBDEs 3073. HEALTH EFFECTS1,000 mg/kg/day during gestation and lactation (Biesemeier et al. 2011; Fujimoto et al. 2011) or doses up to 20 mg/kg/day from PND 2 to 15 (Rice et al. 2007). In females, no exposure-related changes inreproductive development were reported in offspring exposed to decaBDE at doses up to 1,000 mg/kg/day during gestation and lactation (Biesemeier et al. 2011; Fujimoto et al. 2011) or doses up to 20 mg/kg/day from PND 2 to 15 (Rice et al. 2007). In in vitro assays, lower-brominated PBDEs havebeen shown to be mildly estrogenic, anti-androgenic, anti-progestagenic, and anti-glucocorticogenic and higher-brominated PBDEs have been shown to be mildly anti-estrogenic (Hamers et al. 2006; Kojima etal. 2009; Meerts et al. 2001; Mercado-Feliciano and Bigsby 2008a, 2008b; Stoker et al. 2005). PBDEshave also been shown to alter expression and activity of steroidogenic enzymes in vitro (Canton et al. 2005, 2008; He et al. 2008a; Karpeta et al. 2011, 2013; Wang et al. 2011c; Zhao et al. 2011). The developing immune system may also be a target of concern in infants and children exposed toPBDEs. In a one-generation study in rats, F0 rats were exposed to pentaBDE at 0, 0.5, 5, or25 mg/kg/day via gavage for 70 days prior to mating, through mating, gestation, and lactation (PND 21),and PND 43 F1 rats were assessed for serum immunoglobin levels, B and T lymphocyte quantification in the spleen, spleen cell proliferation in vitro, and immune organ weight and histology and PND 56 ratswere assessed for immune function in the KLH antigen immune challenge (Bondy et al. 2013). Observed effects at PND 43 included reduced serum IgE and IgG1 levels in females; a significant, dose-relatedreduction in the proportion of B cells and a significant concomitant increase in the proportion of T cells inthe spleen in males and females; increased proliferation of unstimulated spleen cells harvested from males and females; and histopathological changes in the thymic cortex in males. No exposure-related changes were observed in immune function in PND 56 rats (Bondy et al. 2013). Two studies examined immunefunction in PND 28 mice was assessed using the RSV intranasal infection test following exposure to decaBDE from GD 10 to PND 21 at doses ranging from 3.3 to 3,000 mg/kg/day (Watanabe et al. 2008, 2010b). At doses ≥260 mg/kg/day, RSV pulmonary viral titers were elevated at 1–5 days post-infection, and exposure to 3,100 mg/kg/day exacerbated the histopathological changes in the lung caused by RSVinfection. In the only developmental immune study in humans, a reduced risk of atopic dermatitis wasfound in Japanese infants (diagnosed at 7 months of age by questionnaire) with higher PBDEconcentrations in umbilical cord blood (Ochiai et al. 2014). PBDEs 3083. HEALTH EFFECTS3.8 BIOMARKERS OF EXPOSURE AND EFFECTBiomarkers are broadly defined as indicators signaling events in biologic systems or samples. They have been classified as markers of exposure, markers of effect, and markers of susceptibility (NAS/NRC1989).A biomarker of exposure is a xenobiotic substance or its metabolite(s) or the product of an interactionbetween a xenobiotic agent and some target molecule(s) or cell(s) that is measured within a compartmentof an organism (NAS/NRC 1989). The preferred biomarkers of exposure are generally the substanceitself, substance-specific metabolites in readily obtainable body fluid(s), or excreta. However, severalfactors can confound the use and interpretation of biomarkers of exposure. The body burden of asubstance may be the result of exposures from more than one source. The substance being measured maybe a metabolite of another xenobiotic substance (e.g., high urinary levels of phenol can result fromexposure to several different aromatic compounds). Depending on the properties of the substance (e.g., biologic half-life) and environmental conditions (e.g., duration and route of exposure), the substance and all of its metabolites may have left the body by the time samples can be taken. It may be difficult toidentify individuals exposed to hazardous substances that are commonly found in body tissues and fluids(e.g., essential mineral nutrients such as copper, zinc, and selenium). Biomarkers of exposure to PBDEsare discussed in Section 3.8.1.Biomarkers of effect are defined as any measurable biochemical, physiologic, or other alteration within anorganism that, depending on magnitude, can be recognized as an established or potential health impairment or disease (NAS/NRC 1989). This definition encompasses biochemical or cellular signals oftissue dysfunction (e.g., increased liver enzyme activity or pathologic changes in female genital epithelialcells), as well as physiologic signs of dysfunction such as increased blood pressure or decreased lungcapacity. Note that these markers are not often substance specific. They also may not be directlyadverse, but can indicate potential health impairment (e.g., DNA adducts). Biomarkers of effects causedby PBDEs are discussed in Section 3.8.2.A biomarker of susceptibility is an indicator of an inherent or acquired limitation of an organism’s ability to respond to the challenge of exposure to a specific xenobiotic substance. It can be an intrinsic genetic orother characteristic or a preexisting disease that results in an increase in absorbed dose, a decrease in the biologically effective dose, or a target tissue response. If biomarkers of susceptibility exist, they are discussed in Section 3.10, Populations That Are Unusually Susceptible. PBDEs 3093. HEALTH EFFECTS3.8.1 Biomarkers Used to Identify or Quantify Exposure to PBDEsPBDEs are persistent environmental contaminants that accumulate in adipose tissue, serum, and breastmilk serum of the general population (see Section 3.4, Toxicokinetics). Therefore, PBDE concentrations in these tissues are indicators of general exposure for PBDEs, and PBDE concentrations in maternaladipose tissue, serum, breast milk, and cord serum are useful as markers of maternal body burdens as wellas lactational and in utero exposures (Ahn et al. 2009; CDC 2015; Huwe et al. 2008; Kim et al. 2012a; Marchitti et al. 2013; Mazdai et al. 2003; Meironyté Guvenius et al. 2001). The National Report on Human Exposure to Environmental Chemicals (CDC 2015) has methods for many PBDE congeners to identify exposure and provides reference values representative of U.S. population for comparison with exposures to potentially more exposed populations; it is continually updated on-line (see http://www.cdc.gov/exposurereport/ to look for the most recent U.S. human PBDE exposureinformation). Studies have also proposed that PBDE concentrations in hair are also useful markers ofexposure (Aleksa et al. 2012a, 2012b; Kucharska et al. 2015; Liu et al. 2016; Malarvannan et al. 2013; Poon et al. 2014), although Zheng et al. (2011) reported that PBDE concentrations in hair were notcorrelated with PBDE concentrations in indoor dust from urban, e-waste, and rural areas in South China.It is not clear if the lack of correlation between PBDE concentrations in indoor dust and hair reported byZheng et al. (2011) indicates that hair is a poor monitor of exposure or if exposure via dust was notsignificant for the Chinese subjects involved in the study. Other potential exposure sources, includingoutdoor dust for individuals living near an e-waste area or ingestion of contaminated food, were notcontrolled for in this study. Urinary bromophenols have also been suggested as potential biomarkers ofhuman exposure to PBDEs (Feng et al. 2016a; Ho et al. 2015). Estimates of PBDE serum concentrations among electronics-dismantling workers before and after exposure-free vacation (median duration 28 days, range 21–35 days) indicate that the higher-brominatedcongeners have shorter half-lives than lower-brominated congeners (Sjödin et al. 1999b). The mediansand ranges of percentage decreases in serum concentrations, based on 5–11 measurements per congener,were 14 (range 3.5–39), 14 (2.1–38), 14 (6.7–42), 30 (7.9–52), and 66 (47–100) for BDE 47, BDE 153, BDE 154, BDE 183, and BDE 209, respectively. Although actual half-lives were not calculated, the datasuggest that the half-lives of the lower-brominated congeners were <1 year. Similar findings were found in another study of eight PBDE-exposed workers (four electronics dismantlers and four workers in afactory making flame-retarded rubber), where calculated apparent half-lives were: 15 days for BDE 209;28, 39, and 18 days for nonabrominated congeners BDE 208, BDE 207, and BDE 206; 37, 72, 85, and PBDEs 3103. HEALTH EFFECTS91 days for BDE 203 and three other octabrominated congeners of uncertain chemical structure; and 94 days for BDE 183 (Thuresson et al. 2006).A study in Sweden examined the feasibility of using feces as a noninvasive matrix to determine serumconcentrations of PBDEs in toddlers for biomonitoring purposes (Sahlström et al. 2015). The cohortconsisted of 22 healthy toddlers 11–15 months of age. Twelve (tri- to decabrominated) BDEs were analyzed, but only nine could be quantified; not detected were BDE 28, BDE 99, and BDE 100. Median concentrations of BDEs in feces ranged from 18 ng/g lipid for BDE 209 to 0.055 ng/g lipid for BDE 196.Concentrations of tetra- to octaBDE in serum were significantly higher than in feces; the highestdifference was for BDE 153. BDE 209 was significantly more concentrated in feces than in serum. Significant correlations in concentrations were found for all BDEs detected, except BDE 197 andBDE 203.Congener patterns in humans may provide information on the nature or pathway of PBDE exposures(Hooper and McDonald 2000). Low tetra:deca congener ratios are suggestive of direct, recent, oroccupational exposures to the parent PBDE mixture. Higher ratios may indicate an environmentalpathway where exposures result from PBDEs that have leached from the parent mixtures and have beendegraded in the environment, although they may also reflect metabolic debromination following exposureto parent PBDE mixtures.3.8.2 Biomarkers Used to Characterize Effects Caused by PBDEsBiomarkers of effects for PBDEs are likely to be common to the general class of halogenated aromatic hydrocarbons, rather than specific for PBDEs, because PBBs, PCBs, and other structurally similarchemicals cause generally similar effects.As discussed in Chapter 2, the thyroid, nervous system, and reproductive system are critical targets ofexposure to lower-brominated PBDEs in animals, and evidence from human studies also suggests thatthese systems may be a target of human toxicity. Critical effects used for derivation of the MRLs forlower-brominated PBDEs include: endocrine effects in rats (reduced serum T4) for the intermediate inhalation MRL (Great Lakes Chemical Corporation 2000); endocrine effects in F0 rat dams (reduced serum T4) and reproductive and neurobehavioral effects in F1 adult offspring (impaired spermatogenesis,ultrastructural changes in ovaries, increased resorptions in F1 females mated to unexposed males, andincreased spontaneous motor activity) for the acute oral MRL (Kuriyama et al. 2005, 2007; Talsness et al. PBDEs 3113. HEALTH EFFECTS2005); and reproductive effects in male rats (decreased serum testosterone) for the intermediate oral MRL(Zhang et al. 2013b). There are several potential biomarkers for these effects, such as alterations in serumthyroid or reproductive hormone levels or changes in neurobehavioral screens; however, none of theseeffects are specific to PBDE exposure.Similarly, the nervous system and pancreas are critical targets of exposure to decaBDE in animals. Thereare two human studies suggesting delayed cognitive development in children exposed to decaBDE (Chaoet al. 2011; Gascon et al. 2012) and limited evidence for an association between BDE 153 and diabetes(BDE 209 was not assessed) in humans (Airaksinen et al. 2011; Lee et al. 2011; Lim et al. 2008; Smarr etal. 2016). Critical effects used for derivation of the MRLs for decaBDE include altered neurobehavior in juvenile mice following neonatal exposure for the acute oral MRL (Johansson et al. 2008; Viberg et al. 2003b) and altered insulin homeostasis (elevated serum glucose levels) in rats for the intermediate oral MRL (Zhang et al. 2013a). Again, there are potential biomarkers for these effects, such as alterations inserum glucose levels or changes in neurobehavioral screens; however, none of these effects are specific todecaBDE exposure.3.9 INTERACTIONS WITH OTHER CHEMICALSOnly one study was located that provided information regarding interactions between PBDE and otherchemicals in humans. Fitzgerald et al. (2012) examined the association between exposure to PBDEs and neuropsychological function among 144 adult men and women residents of upper Hudson Rivercommunities and also studied the possible interactive effects of exposure to PCBs. A series of 34 tests toassess cognitive and motor function, affective state, and olfactory function were conducted. Serumsamples were collected and analyzed for concentrations of nine of the most commonly detected PBDEcongeners in human serum (ΣPBDE) and 30 PCB congeners that usually constitute 95% of the congeners found in human serum (ΣPCB). After adjustment for relevant confounders, the results of multiple linearregression analyses showed no significant associations between ΣPBDE and scores on theneuropsychological tests. However, in subjects with a ΣPCB concentration above the median of 467 ppb (on a lipid basis), an increase in ΣPBDE concentration from the 25th to the 75th percentile was significantly associated with decreases between 7 and 12% in scores of some tests of memory and learning. Tests also showed that the interaction was greater than additive. No specific mechanism was proposed for the interaction between PBDEs and PCBs. PBDEs 3123. HEALTH EFFECTSA few animal studies have examined the effects of the interaction of PBDEs and other chemicals onthyroid-related parameters. For example, Hallgren and Darnerud (2002) reported that daily oraladministration of 18 mg/kg/day BDE 47 and 4 mg/kg/day Aroclor 1254 to female rats for 14 daysresulted in a significant decrease in plasma free T4 that suggested additive effects compared toadministration of either compound alone. The reduction in free T4 coincided with a decrease in ex vivobinding of 125I-T4 to the hormone transporter TTR and induction of the microsomal enzymes EROD and MROD, and led the investigators to suggest that the reduced free T4 could be explained by alterations in serum binding to TTR caused by metabolites of Aroclor 1254 or BDE 47. In the same study, co-administration of BDE 47 and a technical mixture of chlorinated paraffins resulted in a reduction in plasma free T4 and induction of EROD that were greater than the effects of the single compounds, indicating synergistic effects. Consistent with the findings of Hallgren and Darnerud (2002), Miller et al. (2012) reported that oral co-administration of equimolar doses of the commercial PBDE mixture DE-71 and a mixture of various Aroclors to pregnant rats from GD 6 until PND 21 significantly reducedcirculating levels of total T4 in male and female offspring from PND 7 to 21 in a manner that indicated additive effects. The reduction in total T4 was dose-dependent over a wide range of doses, 3– 40 µmol/kg/day. The mixture of Aroclors resembled the PCB congener pattern found in contaminated fish consumed by residents near the Fox River in Wisconsin.Wang et al. (2011a) found little evidence of interactions between perfluorooctane sulfonate (PFOS) and BDE 47 on postnatal levels of total T3 and T4 in serum from rats (dams and their offspring) administeredthe chemicals in the diet on GD 1 through PND 14. The investigators also examined the transcriptionaland translational expression of several thyroid hormone-mediated genes in neonates’ brains and found that PFOS and BDE 47 had interactive effects on only levels of brain-derived neurotrophic factor(BDNF); the chemicals showed a synergistic effect on PND 1 in the cortex and an antagonistic effect on PND 14 in the hippocampus. A possible mechanism for the interaction was not discussed. He et al.(2011) also found lack of interactive effects between BDE 47 and 2,2’,4,4’-tetrachlorobiphenyl (PCB 153) on serum T4 levels in rats. In this study, 10-day-old pups received a dose of up to 10 mg/kg BDE 47 and/or 5 mg/kg PCB 153. At 2 months of age, there was no evidence of any interaction between the chemicals on serum levels of T4. However, performance on a Morris water maze was significantlyimpaired in the group dosed with 5 mg/kg BDE 47 and PCB 153 and in the group dosed with 10 mg/kgBDE 47 and PCB 153 relative to the groups dosed with either chemical alone. The effect of the combinedaction of BDE 47 and PCB 153 on performance on a Morris maze had been reported also in an earlierstudy (He et al. 2009). PBDEs 3133. HEALTH EFFECTSTwo studies by Eriksson and coworkers (Eriksson et al. 2006; Fischer et al. 2008), examined theinteractive effects of BDE 99 and 2,2’,5,5’-tetrachlorobiphenyl (PCB 52) or BDE 99 and methyl mercuryon neurodevelopmental parameters in mice. In the first study, combined administration of equimolaramounts of BDE 99 and PCB 52 (1.4 µmol/kg each; 0.8 mg/kg BDE 99, 0.4 mg/kg PCB 52) to 10-dayold male mice resulted in reductions in spontaneous activity and habituation capability at 4 and 6 monthsof age that were significantly more pronounced than those obtained with a single much higher dose(14 µmol/kg) of PCB 52. This led the investigators to suggest that the interaction was greater thanadditive and that different mechanisms may be involved and/or different regions of the brain are differently affected. Using the same protocol, Fischer et al. (2008) reported that coexposure to BDE 99 and methyl mercury exacerbated neurobehavioral defects manifested as alterations in spontaneous behavior, lack of habituation, and impaired learning/memory in male mice tested during the first 6 monthsof life. BDE 99 and methyl mercury also interacted to promote a significant reduction in the density ofcholinergic nicotinic receptors in the hippocampus, suggesting that the neurobehavioral alterations mayinvolve the cholinergic system. The interactive properties of BDE 99 and methyl mercury were alsoassessed in a study in rats (Zhao et al. 2014). Combined administration of these chemicals to pregnantrats during gestation and up to PND 21 resulted in enhanced methylmercury-induced neurotoxicity in thepups compared to treatment with methylmercury alone. Specifically, tests showed delayed appearance ofnegative geotaxis reflexes, impaired motor coordination, and induction of oxidative stress in thecerebellum from pups. While retention of mercury was not affected by co-exposure to BDE 99, combined exposure to methylmercury and BDE 99 seemed to increase the concentrations of BDE 99 in selected brain regions from pups. The investigators concluded that combined exposure to methylmercuryand BDE 99 produced oxidative stress due to inhibition of antioxidant enzymes and production of freeradicals. In a study of the mechanism(s) of neurotoxicity of BDE 47, He et al. (2009) reported that thecombined administration of BDE 47 and PCB 153 to 10-day-old rats induced ultrastructural alterations inneurons in the hippocampal CA1 region, assessed at the age of 2 months, in a manner that suggested asynergistic mode of action. These alterations were associated with changes in expression of mRNA andproteins involved in three apoptosis pathways. For the most part, the changes in expression levels of thevarious factors due to the combined action of BDE 47 and PCB 153 were consistent with an additiveinteraction between the two chemicals.A study in rats reported that gestational and lactational exposure to 5.7 or 11.4 mg/kg/day of thecommercial PBDE mixture DE-71 did not alter cochlear function in adult offspring and neither did exposure to 3 mg/kg/day of an environmental PCB mixture (equimolar to 5.7 mg/kg/day of the PBDEmixture) (Poon et al. 2011). Exposure to 6 mg/kg/day of the PCB mixture did impair cochlear function. PBDEs 3143. HEALTH EFFECTSHowever, exposure of the rats to the combined two low doses of the chemicals, neither of which aloneaffected cochlear function, resulted in a deficit in cochlear function in the offspring similar to that in thehigh-dose PCB group, suggesting an additive effect of the mixture on the outcome measured. Themechanism of the interaction was not elucidated, but it did not appear to be directly related to reductionsin serum T4. The PCB mixture used in this study was the Fox River mixture (see above Miller et al. 2012).It should be noted that the alterations in health outcomes in humans or animals summarized inSection 3.2, Discussion of Health Effects by Route of Exposure, that were caused by exposure tocommercial or environmental PBDE mixtures, are in fact the result of interactions between the individualcomponents of the particular PBDE mixture.3.10 POPULATIONS THAT ARE UNUSUALLY SUSCEPTIBLE A susceptible population will exhibit a different or enhanced response to PBDEs than will most personsexposed to the same level of PBDEs in the environment. Reasons may include genetic makeup, age,health and nutritional status, and exposure to other toxic substances (e.g., cigarette smoke). These parameters result in reduced detoxification or excretion of PBDEs, or compromised function of organsaffected by PBDEs. Populations who are at greater risk due to their unusually high exposure to PBDEsare discussed in Section 6.7, Populations with Potentially High Exposures.Information was located on a small part of the U.S. population that might be unusually susceptible to PBDEs. As indicated in Section 3.4.4.2, the detection of PBDEs in human breast milk samples suggeststhat breast milk represents an elimination route of absorbed PBDEs in women. Both lower and higherbrominated congeners have been detected in breast milk (Antignac et al. 2008, 2009; Malarvannan et al.2013; Park et al. 2011; Schecter et al. 2010). Therefore, women with high body burdens of PBDEs who breastfeed may be placing their infants at a higher risk of potential health effects, although it is unclearthe degree to which PBDEs are cleared from the body during breast feeding (Hooper et al. 2007;Jakobsson et al. 2012; LaKind et al. 2009; Thomsen et al. 2010). In general, however, any risks fromexposures in mother’s milk are outweighed by the benefits of breastfeeding.Pregnant women and developing infants and fetuses should be viewed as possibly sensitive populationsfor exposure to lower-brominated PBDEs, as they are for other thyroid hormone disrupting chemicals(Glinoer 1990; McDonald 2002; Morreale de Escobar et al. 2000). The condition of pregnancy normally PBDEs 3153. HEALTH EFFECTSputs a significant strain on the maternal thyroid system, which can be exacerbated by iodine deficiency;according to data from 1988 to 1994, iodine deficiency is prevalent in approximately 12% of the generalpopulation and 15% of women of child-bearing age in the United States (Hollowell et al. 1998). Thyroid hormones are essential for normal development of the nervous system, lung, skeletal muscle, and possiblyother organ systems, and the fetus is dependent on maternal thyroid hormones at least until the fetal thyroid begins to produce T4 and T3, which occurs in humans at approximately 16–20 weeks of gestation(Zoeller and Crofton 2000). As discussed in Section 3.2.2.2 (Endocrine Effects subsection), human studies have found inconsistent associations between PBDE levels in maternal/infant serum, cord blood, or breast milk and thyroid hormone levels in pregnant women and infants. However, decreased maternal, fetal, and neonatal serum T4 levels have been consistently reported in animals exposed to lowerbrominated PBDEs (but not decaBDE). Therefore, it is unclear whether or not exposure to lowerbrominate PBDEs will cause thyroid hormone disruption in humans, and mechanistic differences mayaccount for the observed interspecies differences (as discussed in Section 3.5.3, Animal-to-HumanExtrapolations).Developing fetuses, infants, and children should also be regarded as a possibly sensitive population withregards to neurodevelopmental effects. As discussed in Section 3.7 (Children’s Susceptibility), numerousstudies have reported results suggestive of an effect of PBDE on neurodevelopment in children. PBDElevels in cord blood, maternal or infant serum, and/or breast milk have been correlated with cognitivedeficits (including impaired verbal memory and comprehension), adaptive behavior deficits, increased impulsivity and impaired attention, poor social competence, and impaired fine motor coordination in infants and children. Additionally, pre- and peri-natal studies in animals consistently report neurodevelopmental effects following exposure to lower-brominated PBDEs and decaBDE (at higherdoses). The susceptibility of the developing nervous system may be due to neuroendocrine effects, as discussed in Section 3.6 (Toxicities Mediated Through The Neuroendocrine Axis). For example, neuronal migration and differentiation of fetal hNPCs were significantly impaired following in vitro exposure to tetraBDE (Schreiber et al. 2010), and decaBDE was shown to be a thyroid hormone receptorantagonist that significantly inhibited T4-induced dendritic arborization in cultured rat cerebellar Purkinjecells (Ibhazehiebo et al. 2011).People with exposure to anti-thyroid drugs (e.g., lithium), thyroid disease, or otherwise compromised thyroid function might have a more pronounced response to PBDEs because of their underlyinglimitations in thyroid hormone production. Similarly, people with compromised function of other organs, PBDEs 3163. HEALTH EFFECTSsuch as those with liver or kidney diseases (e.g., liver cirrhosis or hepatitis), could be considered moresusceptible to health effects of PBDEs.3.11 METHODS FOR REDUCING TOXIC EFFECTSThis section will describe clinical practice and research concerning methods for reducing toxic effects ofexposure to PBDEs. However, because some of the treatments discussed may be experimental andunproven, this section should not be used as a guide for treatment of exposures to PBDEs. When specificexposures have occurred, poison control centers and medical toxicologists should be consulted formedical advice. No texts were located that provide specific information about treatment followingexposures to PBDEs; however, recommendations based on experiences with PCBs are relevant. Thefollowing texts provide specific information about treatment following exposures to PCBs:Caravati EM, Mcguigan MA, Whyte IM, et al. 2004. Polyhalogenated biphenyls. In: Dart RC, ed. Medical toxicology. 3rd ed. Philadelphia, PA: Lippincott Williams & Wilkins, 1342-1343.Leikin JB, Pauloucek FP. 2008. Polychlorinated biphenyls. In: Leikin JB, Pauloucek FP, eds. Poisoning and toxicology handbook. 4th ed. Boca Raton, FL: CRC Press, Taylor & Francis Group, 840.The treatment methods discussed below are general methods that would apply to any persistent, lipophilicchemical, and have not been tested for efficacy, indicating that they might not be effective in reducing the toxic effects of PBDEs. There is no indication of hazards associated with the treatments. The methods are particularly appropriate for trying under conditions of acute exposure, but PBDEs are not acutely toxicchemicals. Scenarios where life-threatening acute exposure would occur are unlikely, although accidentalor intentional ingestion of the commercial products is a conceivable concern. The relevance of the methods to common background environmental exposures to these chemicals is unclear, and it isquestionable whether current exposure and tissue levels in the general population are a health concern.3.11.1 Reducing Peak Absorption Following Exposure Ingested PBDEs are absorbed by the gastrointestinal tract of humans and animals (see Section 3.4). Although there are no specific recommendations for clinical treatment of acute intoxication from ingested PBDEs, recommendations based on experiences with PCBs are relevant. Treatments for acute poisonings from PCBs and related substances include the induction of emesis or gastric lavage and stomach pumpingto decrease gastrointestinal absorption of the chemicals (Lemesh 1992). These procedures would not bebeneficial if performed too long after exposure occurred. Administration of activated charcoal as a slurry, PBDEs 3173. HEALTH EFFECTSeither aqueous or mixed with a saline cathartic or sorbitol, is frequently recommended to decrease the gastrointestinal absorption of PCBs, but the value of this treatment for reducing absorption of PCBs,PBBs, and PBDEs is unknown (HSDB 2012). Repetitive administration of activated charcoal might be useful in preventing reabsorption of metabolites. Rice bran fiber decreased absorption of PCBs in the gastrointestinal tract and had a stimulatory effect on fecal excretion of PCBs in rats (Takenaka and Takahashi 1991), but it is unclear if rice bran would be of benefit in poisoned humans.The detection of PBDEs in the serum and fat of people who were occupationally exposed to thesechemicals indicates that PBDEs can be absorbed by the lungs, skin, and/or orally by hand-to-mouth contact. Although no specific methods to reduce absorption of dermally applied or inhaled PBDEs werelocated, multiple washings of contaminated skin with soap and water immediately following exposurehave been suggested to reduce the dermal absorption of PCBs (HSDB 2012). Studies with monkeysshowed that soap and water was as effective as or better than such solvents as ethanol, mineral oil, ortrichlorobenzene in removing PCBs from skin (Wester et al. 1990). Personal protective equipment(e.g., long sleeves, gloves, safety glasses, respiratory protection) and industrial hygiene programs generally help to limit occupational exposures. 3.11.2 Reducing Body BurdenAs discussed in Section 3.4, while decaBDE is absorbed to a lesser degree than lower-brominated PBDEs, all PBDE congeners can accumulate in lipid-rich tissues. However, lower-brominated PBDEs are more likely to accumulate as they are more slowly metabolized and eliminated from the body. Nostudies evaluating methods to reduce body burden of PBDEs were located.Several methods to enhance the elimination of PBBs from the body have been examined in animals and may be applicable to PBDEs. Methods for increasing the elimination of these chemicals include therestriction of caloric intake (to reduce total body fat), and the administration of various agents that interactwith bile acids including activated charcoal, mineral oil, and bile-binding resins such as cholestyramine(Kimbrough et al. 1980; McConnell et al. 1980; Polin and Leavitt 1984; Polin et al. 1985, 1991; Rozman et al. 1982). It should be mentioned, however, that based on the pharmacokinetic considerationsdiscussed in Section 3.8.1, a rapid breakdown of fat, as might occur in dieting, might lead to a transient increase in PBDE levels in serum and other body tissues, possibly posing a significant re-exposureproblem. Although some of the studies observed no enhanced elimination (Kimbrough et al. 1980;McConnell et al. 1980), others identified treatments that were effective in enhancing the biliary and PBDEs 3183. HEALTH EFFECTSintestinal elimination of PBB residues (Polin et al. 1991; Rozman et al. 1982). Polin et al. (1991) found that dietary intervention to reduce PBBs was dose dependent; treatment with 10% mineral oil and a 45%reduction in food intake resulted in a 69 and 23% reduction in body burden in rats fed PBBs at dietaryconcentrations of 0.1 and 100 ppm, respectively (Polin et al. 1991). A combination of mineral oil, colestipol, and dietary restriction was successful in reducing the PBB body burdens in chickens (Polin and Leavitt 1984; Polin et al. 1985), while each treatment alone had no effect in reducing PBB bodyburden. A 3-week treatment regimen that included dietary supplements of polyunsaturated oil, vitamins, and minerals, and heat stress has been applied in a pilot study to seven human subjects that were known tohave been exposed to PBBs; following treatment, statistically significant reductions were measured inPBB concentrations in fat (Schnare et al. 1984). Although the lack of a separate control group complicates interpretation of the results of this study (each subject served as his/her own control), this treatment was developed for the purpose of reducing body burdens of fat-soluble psychoactive drugs(Schnare et al. 1984).Additionally, a few human studies that have evaluated methods to reduce body burden of PCBs, anotherclass of lipophilic compounds, may be applicable to PBDEs. A liquid diet was used for 16 individualswho developed symptoms following exposure to PCBs and polychlorinated dibenzofurans (Imamura and Tung 1984). Symptoms were reduced several months after the fasting period. This study is limited in that a control group was not used, and body burdens were not measured. Based on information for PCBs, mobilization of PBDEs from adipose tissue is not recommended in individuals with hepatic or renaldisease (Lemesh 1992). More recently, 14 individuals administered 15 g/day of dietary olestra (a nonabsorbable lipid in potato crisps) for 1 year showed a steady decline in serum lipid concentrations ofPCBs (Jandacek et al. 2014). At the end of 1 year, PCB concentrations were significantly decreased by~8% compared with pre-trial values. However, PCB concentrations were not significantly decreasedcompared with 14 concurrent controls administered 15 g/day of dietary olive oil (which showed anonsignificant ~4% decrease in PCB concentration compared with pre-trial values). Further studies needto be conducted with larger study groups to determine the efficacy of olestra for reducing body burden ofPCBs and/or other lipophilic compounds. 3.11.3 Interfering with the Mechanism of Action for Toxic EffectsThere are no known methods for interfering with the mechanism of action of PBDEs. Although themechanism of action of PBDEs is not completely understood, PBDEs share some toxicological properties with other structurally similar polyhalogenated aromatic compounds, particularly PBBs, PCBs, CDDs, PBDEs 3193. HEALTH EFFECTSand CDFs (ATSDR 1994, 1998, 2000). Experimental evidence indicates that some PBB congeners exerttoxic actions by a process involving several steps (Safe 1984). This process begins with the binding ofparticular congeners to the AhR and leads to enhancement of the CYP1A1 gene expression (see Section 3.5). The toxicity of specific congeners is related to their ability to assume a coplanarconfiguration for binding to AhR (Hardy 2002a). It may be inferred that interfering with the initial step,binding to the receptor, or with any of the subsequent steps, would possibly prevent the expression of thetoxic effects. Several compounds have been identified that partially antagonize one or more AhRmediated responses (Bannister et al. 1989); their use, however, has been limited to experimental studies in animals. These compounds were successful antagonists when given before or at the same time as thepotent AhR activator TCDD (Bannister et al. 1989). In contrast to PBBs, PCBs, and related compounds, AhR binding affinity of PBDE congeners is not correlated with planarity (Chen et al. 2001). This may bedue, in part, to increased distance between the phenyl rings relative to PBBs and PCBs. It has also beenspeculated that the large size of the bromine atoms of PBDEs relative to chlorine atoms of PCBs maydistort the AhR binding site so that coplanar configuration is not required (Chen et al. 2001). Because AhR binding by PBDEs apparently differs in some respects from AhR binding by PBBs and other related compounds, AhR antagonists identified via experiments with PBBs, TCDD, and related compoundsmight not effectively antagonize AhR binding and effects of PBDEs.PBDEs may also cause toxicity by other mechanisms of action. PBDE-induced decreases in thyroid T4 hormone, which can affect neurobehavioral development, are likely to involve multiple mechanisms (seeSection 3.5.3). These include induction of hepatic microsomal enzymes, particularly UDPGT, which can increase the rate of T4 conjugation and excretion, and metabolic formation of hydroxy-metabolites ofPBDEs. PBDEs and their hydroxy metabolites can bind with high affinity to thyroid transport proteinsbecause they are structurally similar to T4 hormone (i.e., are also hydroxy-halogenated diphenyl ethers)(see Section 3.5.2). Effects of PBDEs on thyroid status via induction of hepatic enzymes, however, areunlikely to occur in humans, and the impact of hydroxy-metabolites on serum T4 needs furtherclarification. Effects of PBDEs on the function and development of the nervous system could alsoinvolve disruption of calcium homeostatic mechanisms and intracellular signalling events (Chen et al.2010; Dingemans et al. 2008, 2010a, 2010b; Fan et al. 2010; Kodavanti 2003; Kodavanti and Derr-Yellin2001, 2002; Smolnikar et al. 2001; Wiegand et al. 2001), altered cholinergic or dopaminergic functions(Ankarberg et al. 2001; Bradner et al. 2013; Dreiem et al. 2010; Fischer et al. 2008; Mariussen and Fonnum 2002, 2003; Mariussen et al. 2003; Slotkin et al. 2013; Viberg and Eriksson 2011; Viberg et al.2002, 2003a, 2004b, 2005), and/or free radical-induced neuronal death (Chen et al. 2010; He et al. 2008b; PBDEs 3203. HEALTH EFFECTSHuang et al. 2010; Reistad et al. 2002). Clinical interventions designed to interfere with theaforementioned mechanisms have yet to be developed.3.12 ADEQUACY OF THE DATABASE Section 104(I)(5) of CERCLA, as amended, directs the Administrator of ATSDR (in consultation with theAdministrator of EPA and agencies and programs of the Public Health Service) to assess whetheradequate information on the health effects of PBDEs is available. Where adequate information is notavailable, ATSDR, in conjunction with the National Toxicology Program (NTP), is required to assure theinitiation of a program of research designed to determine the health effects (and techniques for developingmethods to determine such health effects) of PBDEs.The following categories of possible data needs have been identified by a joint team of scientists fromATSDR, NTP, and EPA. They are defined as substance-specific informational needs that if met would reduce the uncertainties of human health assessment. This definition should not be interpreted to mean that all data needs discussed in this section must be filled. In the future, the identified data needs will be evaluated and prioritized, and a substance-specific research agenda will be proposed.3.12.1 Existing Information on Health Effects of PBDEsThe existing data on health effects of inhalation, oral, and dermal exposure of humans and animals toPBDEs are summarized in Figure 3-8. The purpose of this figure is to illustrate the existing informationconcerning the health effects of PBDEs. Each dot in the figure indicates that one or more studies provideinformation associated with that particular effect. The dot does not necessarily imply anything about thequality of the study or studies, nor should missing information in this figure be interpreted as a “dataneed”. A data need, as defined in ATSDR’s Decision Guide for Identifying Substance-Specific Data Needs Related to Toxicological Profiles (ATSDR 1989), is substance-specific information necessary toconduct comprehensive public health assessments. Generally, ATSDR defines a data gap more broadlyas any substance-specific information missing from the scientific literature.Studies on the systemic and carcinogenic effects of PBDEs in humans are based primarily on tissue PBDE levels without knowledge of route of exposure; in most cases exposure was attributed to the oralroute (Figure 3-8). Information on health effects of PBDEs in animals is available for all effectcategories, but is mainly limited to oral exposure studies in animals. PBDEs 3213. HEALTH EFFECTSFigure 3-8. Existing Information on Health Effects of Polybrominated DiphenylEthers (PBDEs)Inhalation Oral Dermal Death Acute IntermediateChronic Immunologic/LymphoreticNeurologicReproductiveDevelopmentalGenotoxicCancer Systemic Human Death Acute IntermediateChronic Immunologic/LymphoreticNeurologicReproductiveDevelopmentalGenotoxicCancer Systemic Inhalation Oral Dermal Animal Existing Studies PBDEs 3223. HEALTH EFFECTS3.12.2 Identification of Data Needs Acute-Duration Exposure. Acute-duration studies of lower-brominated PBDEs have documented effects mainly on the developing nervous system, the developing reproductive system, the thyroid, and the liver of orally exposed rats and mice (Bruchajzer 2011; Bruchajzer et al. 2010, 2011; Darnerud and Sinjari 1996; Dingemans et al. 2007; Dufault et al. 2005; Eriksson et al. 2001, 2002b, 2006; Fischer et al. 2008; Fowles et al. 1994; Gee and Moser 2008; Hallgren and Darnerud 2002; Hallgren et al. 2001, 2015; He et al. 2009, 2011; Hoppe and Carey 2007; Kuriyama et al. 2005, 2007; Richardson et al. 2008; Sand etal. 2004; Stoker et al. 2004; Talsness et al. 2005, 2008; Viberg et al. 2002, 2003a, 2004a, 2004b, 2005, 2006; Zhou et al. 2001, 2002). The most sensitive effects were observed in F0 and F1 rats exposed to ≥0.06 mg/kg of pentaBDE on GD 6, including decreased maternal serum T4 in dams, decreased number ofspermatids and sperm and daily sperm production in F1 males, decreased relative epididymis weight in F1 males, altered neurobehavior in F1 males, ultrastructural changes in ovaries of F1 females, andincreased resorptions in F1 females mated to unexposed males (Kuriyama et al. 2005, 2007; Talsness etal. 2005). Collectively, these end points were selected as a basis for an acute MRL for lower-brominated PBDEs. While hepatic effects were consistently observed, they occurred at much higher doses (≥8 mg/kg/day). Two studies in mice indicate that immunosuppression is a potentially critical health endpoint for acute exposure to lower-brominated PBDEs; additional studies may provide more support forthis end point (Darnerud and Thuvander 1998; Fowles et al. 1994; see discussions of data needs forImmunotoxicity). Several acute-duration studies of decaBDE have also documented effects on the developing nervoussystem in rats (Chen et al. 2014; Viberg et al. 2007) and mice (Buratovic et al. 2014; Johansson et al. 2008; Rice et al. 2007, 2009; Viberg et al. 2003b). The most sensitive neurobehavioral effects, decreased open field activity and impaired habituation, were observed in 2–6-month-old male mice that were exposed once to decaBDE at doses ≥2.22 mg/kg on PND 3 (Buratovic et al. 2014; Johansson et al. 2008;Viberg et al. 2003b). This effect was not observed at 1.34 mg/kg (Buratovic et al. 2014; Johansson et al.2008). These neurobehavioral end points were selected as the basis for an acute MRL for decaBDE. The remaining acute database for decaBDE provides only limited data regarding hepatic, endocrine, bodyweight, and developmental effects in rats and mice (Bruchajzer et al. 2010; Carlson 1980b; Chi et al.2011; IRDC 1974; NTP 1986; Sakamoto et al. 2013; Zhou et al. 2001). Additional studies, specificallyneurodevelopmental studies in species other than mice and immunotoxicity screens, would help to clearly PBDEs 3233. HEALTH EFFECTSestablish the most sensitive target and species for acute exposure, as well as which animal toxicity data are the most relevant to humans and useful for assessing acute health risks of decaBDE.The inhalation database for acute-duration exposure to lower-brominated PBDEs is essentially limited toa single 14-day unpublished industry-sponsored studies of octaBDE in rats (Great Lakes ChemicalCorporation 1978). A liver effect level was identified in this study, but MRL estimation is precluded bysmall animal groups and incomplete assessment of other toxicological end points, particularly a lack ofinformation on thyroid hormone levels. Additional dose-response studies may provide an adequate basis for derivation of an acute inhalation MRL for lower-brominated PBDEs. Acute-duration inhalation exposure toxicity studies of decaBDE were not located.Intermediate-Duration Exposure. Available intermediate-duration oral studies in animals indicatethat the male reproductive system, the developing nervous system, the thyroid, and the liver are the mainsystemic targets of repeated exposures to lower-brominated PBDEs. The most sensitive end points werereproductive effects in male rats exposed to tetraBDE for 8 weeks, which identified a minimal LOAEL of0.001 mg/kg/day for a 34% decrease in serum testosterone levels (Zhang et al. 2013b). Histological changes in the testes were also observed at ≥0.03 mg/kg/day in the study by Zhang et al. (2013b) and asimilar study in rats by Huang et al. (2015). The minimal LOAEL for decreased serum testosterone was selected as the basis for an intermediate-duration MRL for lower-brominated PBDEs. A study in micesupports that the male reproductive system is a target for tetraBDE toxicity, reporting histologicalchanges in the testes at tetraBDE doses ≥0.045 mg/kg/day for 30 days (Wang et al. 2013). As observed with acute exposure, altered neurobehavior was consistently observed in animals exposed to lowerbrominated PBDEs for intermediate durations during development, with effects occurring at or above thedoses of observed male reproductive effects (≥2 and ≥0.03 mg/kg/day in rats and mice, respectively)(Blanco et al. 2013; Bowers et al. 2015; Branchi et al. 2005; Cheng et al. 2009; Koenig et al. 2012; Ta etal. 2011; Woods et al. 2012). Liver effects, including increased liver weight, hypertrophy, and histopathological changes, were observed following exposure to lower-brominated PBDEs at doses as low as 0.45 mg/kg/day (Becker et al. 2012; Bondy et al. 2011, 2013; Bruchajzer 2011; Bruchajzer et al. 2012; Dunnick et al. 2012; IRDC 1976, 1977; Maranghi et al. 2013; Oberg et al. 2010; WIL Research Laboratories 1984; Zhang et al. 2014, 2015a, 2015b).The thyroid is also a critical target in both adult and developing animals, with consistent observations ofreduced serum T4 levels at doses as low as 2.85 mg/kg/day and enlargement and histological alterations to the thyroid at doses as low as 0.45 mg/kg/day (Bansal et al. 2014; Becker et al. 2012; Bondy et al. 2011, PBDEs 3243. HEALTH EFFECTS2013; Bowers et al. 2015; Driscoll et al. 2009; Dunnick et al. 2012; Ellis-Hutchings et al. 2006; Ernest etal. 2012; Hoppe and Carey 2007; IRDC 1977; Kodavanti et al. 2010; Maranghi et al. 2013; Miller et al. 2012; Poon et al. 2011; Skarman et al. 2005; Stoker et al. 2004, 2005; Szabo et al. 2009; Wang et al.2011a; WIL Research Laboratories 1984; Zhou et al. 2002). Essentially all of the available data onthyroid effects of lower-brominated PBDEs have been obtained from oral studies in rats. It is speculatedthat the extent that PBDEs affect circulating levels of thyroid T4 or T3 might vary with species, and rats are often regarded as more sensitive than humans. Studies designed to elucidate the mechanism(s) ofaction for thyroid and other effects of lower-brominated PBDEs would help to better understand how theanimal toxicity data can best be used to identify target end points and assess health risks in humans.Available intermediate-duration oral studies provide limited evidence for effects in the pancreas, nervoussystem, liver, immune system, reproductive system, and several organ systems in developing animals following repeated exposures to decaBDE. The most sensitive end points were pancreatic effects from a study in male rats exposed to decaBDE for 8 weeks, which identified a minimal LOAEL of0.05 mg/kg/day based on a 12% increase in serum glucose levels (Zhang et al. 2013a). The increase inserum glucose is considered to be part of a spectrum of effects indicative of altered insulin homeostasisand toxicity to the pancreas, including decreased serum insulin and morphological changes in pancreatic islet cells, following decaBDE exposure to doses ≥1 mg/kg/day (Zhang et al. 2013a). This minimal LOAEL for elevated serum glucose levels was used as a basis for the intermediate-duration oral MRL fordecaBDE. Other effects observed following intermediate-duration exposure to decaBDE were observed at higher doses. In animals exposed during development, adverse effects included histopathological changes in the liver, kidney, and testes of rats exposed to ≥2 mg/kg/day (Fujimoto et al. 2011; Tseng et al.2008, 2013); altered hippocampal electrophysiology in rats exposed to 20.1 mg/kg/day (Xing et al. 2009);and impaired immunity in mice exposed to 2,900 mg/kg/day (Watanabe et al. 2010b). In adult animals, observed effects following intermediate-duration decaBDE exposure included decreased anxiety in miceat doses 20 mg/kg/day (Heredia et al. 2012); histopathological changes in the ovaries, liver, spleen, andthymus and altered immune end points (T-cell distribution, lymphocyte proliferation, serumimmunoglobulins) in rats exposed to 300 mg/kg/day (Liu et al. 2012); altered CD4 T-cell function in mice exposed to 800 mg/kg/day (Feng et al. 2016b); reduced serum testosterone, reduced sperm count and viability, and degenerative changes in the seminiferous tubules in mice exposed to 950 mg/kg/day (Sarkaret al. 2015); and histopathological changes in the liver in mice exposed to 9,400 mg/kg/day (Lee et al.2010; Liu et al. 2012; Sakamoto et al. 2013). Since the majority of the observed effects lack supportingevidence from other studies, and many were single-dose studies, additional intermediate-duration studiesevaluating these end points following decaBDE exposure at multiple doses would help to better PBDEs 3253. HEALTH EFFECTSunderstand how the animal toxicity data can best be used to identify target end points and assess health risks in humans. Of particular interest would be additional studies evaluating altered insulin homeostasis,the critical effect that serves as the basis for the current intermediate-duration oral MRL for decaBDE.The inhalation database for intermediate-duration exposure to PBDEs consists of one well-conducted13-week unpublished industry study of octaBDE in rats (Great Lakes Chemical Corporation 2000). Hepatic, nasal, lung, thyroid, and ovarian effects were observed, and a NOAEL for changes in thyroid hormone levels was used as the basis for estimation of an intermediate-duration inhalation MRL. Intermediate-duration inhalation exposure toxicity studies of decaBDE were not located. Chronic-Duration Exposure and Cancer. One chronic study of high purity decaBDE has been conducted. In this study, a commercial decaBDE product (94–97% pure) was fed to rats and mice for103 weeks (NTP 1986). Comprehensive gross and histological examinations were performed on allanimals, but no hematology, clinical chemistry, or urine indices, or thyroid hormone levels, wereevaluated. The lowest tested dose in the study, 1,120 mg/kg/day in male rats, was a LOAEL for a liverlesion (neoplastic nodules) that is precancerous and associated with thrombosis in the same tissue, precluding estimation of an MRL. Additional chronic dose-response information would provideinformation on the NOAEL/LOAEL threshold and an appropriate basis for derivation of a chronic MRLfor decaBDE. Neoplastic effects in this study included increased incidences of neoplastic nodules in theliver in the male and female rats and hepatocellular adenoma or carcinoma (combined) in the male mice.Slightly elevated incidences of thyroid gland follicular cell tumors were additionally observed in exposed male mice, although the increases were equivocal.In the only other chronic oral study available, rats were fed a 77.4% pure commercial decaBDE mixture(containing 21.8% nonaBDE and 0.8% octaBDE) for approximately 2 years (Kociba et al. 1975; Norris etal. 1975a). Evaluations that included clinical signs, body weight, food consumption, hematology, clinicalchemistry, urine indices, and comprehensive histological examinations showed no exposure-relatedeffects. The highest tested dose (1 mg/kg/day) was a NOAEL, but this effect level is not an appropriate basis for MRL estimation due to insufficient sensitivity of the study. In particular, a chronic oral MRLbased on this study would be higher than the intermediate MRL. No exposure-related neoplastic changes were found, but the power of this study to detect carcinogenic effects is limited by the low dose levels.Considering the limitations of the available data, well-designed chronic toxicity studies of lowerbrominated PBDEs may provide adequate bases for MRL derivation and cancer assessment for lowerbrominated PBDEs. Evaluations that include the thyroid and neurobehavioral end points would be PBDEs 3263. HEALTH EFFECTSparticularly informative because acute and intermediate-duration oral studies indicate that the thyroid and developing central nervous system are particularly sensitive targets for lower-brominated PBDEs.Genotoxicity. Limited information was located regarding the genotoxicity of PBDEs in humans. Astudy of Chinese workers exposed to PBDEs suggested an association between levels of total PBDEs in blood and frequency of micronuclei in peripheral lymphocytes, but no causality was demonstrated (Yuanet al. 2008). Moreover, other pollutants could have played a role. A study of the general population fromKorea did not find an association between serum levels of 2,2’,4,4’-tetraBDE and 2,2’,4,4’,5-pentaBDEand chromosome telomere length in peripheral lymphocytes (Shin et al. 2010). Cytogenetic examination of additional populations previously or currently exposed to PBDEs would provide valuable information. A study in mice reported that maternal exposure to decaBDE during gestation resulted in sperm DNAdamage in male offspring examined at 71 days of age (Tseng et al. 2013). Additional studies trying to replicate these results would be helpful. Testing the offspring for fertility would also be informative.DecaBDE and 2,2’,4,4’,5-pentaBDE have been tested in prokaryotic organisms and both yielded negativeresults in gene mutation tests (Evandri et al. 2003; NTP 1986). Relatively few PBDE congeners have been examined for genetic effects in mammalian cells. DNA damage and increased recombination activity have been reported (He et al. 2008a, 2008b; Helladay et al. 1999; Ji et al. 2011; Pellacani et al. 2012). Further studies with the PBDE congeners that are most frequently found in the environment and in human blood and tissues would be valuable. Studies designed to explore possible mechanisms ofgenotoxicity of PBDEs and also of metabolites would also be valuable.Reproductive Toxicity. Reports of five one-generation studies in rats and mice were located. Noexposure-related changes were observed in reproductive end points (number of pregnancies, gestationlength, number, size, or sex ratio of litters) at pentaBDE doses up to 25 mg/kg/day or tetraBDE doses up to 1 mg/kg/day (Bondy et al. 2011, 2013; Koenig et al. 2012; Poon et al. 2011; Ta et al. 2011; Woods etal. 2012). No fertility impairments were observed in F1 males or females that were exposed once topentaBDE on GD 6 at doses up to 0.3 mg/kg/day, when mated to unexposed animals (Kuriyama et al. 2005; Talsness et al. 2005, 2008). However, in a one-generation study in mink, females exposed to pentaBDE at doses ≥0.25 mg/kg/day from pre-mating day 28 through PNW 6 did not whelp (Bull et al. 2007; Zhang et al. 2009). It is not clear in the study reported by Bull et al. (2007) whether mink exposed to 0.25 mg/kg/day never became pregnant or had complete litter loss. However, Zhang et al. (2009)reported that female mink exposed to 0.31 mg/kg/day had no exposure-related changes in mating success;rather, sows showed complete litter loss with 70% showing clear postimplantation loss. Despite a lack ofeffects from most one-generation studies, reduced serum testosterone and testicular damage (increased PBDEs 3273. HEALTH EFFECTSmultinucleated giant cells, germ cell loss, and apoptotic cells) was reported in adult male rats and mice exposed to tetraBDE doses as low as 0.001 and 0.045 mg/kg/day, respectively (Huang et al. 2015; Wanget al. 2013; Zhang et al. 2013b). Reduced serum testosterone in rats exposed to ≥0.001 mg/kg/day for 8 weeks was selected as the critical effect for the intermediate-duration MRL for lower-brominated PBDEs; therefore, additional studies with other congeners evaluating testicular damage may bettercharacterize the reproductive toxic potential of PBDEs and assure the adequacy of the intermediate oralMRL. Also, studies in other species (e.g., rabbits) may be warranted to investigate species-specific reproductive effects and two-generation studies designed to assess effects on fertility in both sexes wouldbetter characterize the potential for reproductive toxicity as a result of exposure to lower-brominatedPBDEs.For decaBDE, information on the reproductive toxicity is limited to a single one-generation oral study in rats that found no exposure-related functional effects following exposure to an impure decaBDE mixture containing lower-brominated PBDEs (77% decaBDE, 22% nonaBDE, 0.8% octaBDE) at dietary doses up to 100 mg/kg/day for 60 days prior to mating through PND 21 (Dow Chemical Co. 1975; Norris et al.1975a). Studies that evaluated reproductive organ histology following exposure to decaBDE generallydid not report detectable effects. As suggested for lower-brominated PBDEs, studies in other species maybe warranted to investigate species-specific reproductive effects and two-generation studies designed to assess effects on fertility in both sexes would better characterize the potential for reproductive toxicity as a result of exposure to decaBDE.Developmental Toxicity. Numerous oral developmental toxicity studies have shown no evidence ofteratogenicity in rats or rabbits exposed to lower-brominated PBDEs or decaBDE, although fetotoxiceffects, including skeletal ossification variations at maternally toxic doses, have occurred with exposuresto lower-brominated mixtures (Argus Research Laboratories 1985a, 1985b; Biesemeier et al. 2011; Blanco et al. 2012; Breslin et al. 1989; Chi et al. 2011; Dow Chemical Co. 1975, 1985; Ellis-Hutchings etal. 2009; Hardy et al. 2002; Life Science Research Israel Ltd. 1987; Norris et al. 1975a; WIL ResearchLaboratories 1986). The available evidence adequately shows that teratogenicity and fetal toxicity is nota critical effect of concern for either lower-brominated PBDEs or decaBDE. However, there is evidencethat the developing nervous and endocrine systems, and potentially the developing reproductive and immune systems are sensitive targets of particular PBDE congeners. Numerous studies show thatdevelopmental exposure can lead to neurological changes at later life stages, as rats and mice exposed pre- or peri-natally to lower-brominated PBDEs or decaBDE show neurobehavioral alterations followingexposure to lower-brominated PBDEs and decaBDE at doses ≥0.06 and ≥2.22 mg/kg/day, respectively, as PBDEs 3283. HEALTH EFFECTSwell as several other neurological effects at higher doses, including delayed ontogeny of reflexes, ultrastructural changes, altered nicotinic receptor density, altered electrophysiology, and altered gene andprotein expression levels (Biesemeier et al. 2011; Blanco et al. 2013; Branchi et al. 2001, 2002, 2005;Cheng et al. 2009; Eriksson et al. 2001, 2002b, 2006; Fischer et al. 2008; Fujimoto et al. 2011; Gee and Moser 2008; He et al. 2009, 2011; Johansson et al. 2008; Koenig et al. 2012; Kuriyama et al. 2004, 2005;Rice et al. 2007; Sand et al. 2004; Ta et al. 2011; Viberg et al. 2002, 2003a, 2003b, 2004a, 2004b, 2005, 2006, 2007; Woods et al. 2012; Xing et al. 2009). Developing animals exposed to lower-brominatedPBDEs, but not decaBDE, also consistently showed decreased serum T4 levels, although these effects may be transient (Blanco et al. 2013; Bondy et al. 2011, 2013; Ellis-Hutchings et al. 2006; Kodavanti etal. 2010; Kuriyama et al. 2007; Miller et al. 2012; Poon et al. 2011; Shah et al. 2011; Skarman et al. 2005;Szabo et al. 2009; Wang et al. 2011a; Zhang et al. 2009; Zhou et al. 2002). There is limited evidence thatdevelopmental exposure to lower-brominated PBDEs may effect reproductive system development(Kuriyama et al. 2005; Talsness et al. 2005, 2008) and that both lower-brominated PBDEs and decaBDEmay effect immune system development (Bondy et al. 2013; Watanabe et al. 2008, 2010b). Additionalstudies evaluating reproductive and immune system development following exposure to PBDEs would better characterize the developmental toxic potential of PBDEs on these systems.Immunotoxicity. Information regarding the immunosuppressive potential of PBDEs is limited.Immune function assays in adult animals are limited to lower-brominated PBDEs. Acute-duration oralstudies in animals exposed to pentaBDE reported that plaque-forming splenic cell antibody response to injected sheep red blood cells was significantly reduced in mice exposed to 72 mg/kg/day pentaBDE for14 days (Fowles et al. 1994) and in vitro production of IgG immunoglobulin from pokeweed mitogenstimulated splenocytes was reduced in mice, but not in rats, exposed to 36 mg/kg/day pentaBDE for14 days (Darnerud and Thuvander 1998). The only intermediate-duration immune function assayreported no exposure-related immune effects in the KLH antibody production assay or the PHA skin challenge in mink exposed to pentaBDE at doses up to 0.78 mg/kg/day for 9 weeks (Martin et al. 2007). The majority of studies report no histological changes in immune tissues with acute- or intermediate-duration exposure to PBDEs. However, as discussed in the Developmental section above, exposure to lower-brominated PBDEs or decaBDE during development may lead to impaired immune function (Bondy et al. 2013; Watanabe et al. 2008, 2010b). Additional oral studies using a battery ofimmunological tests and a lower range of doses for several congeners, including decaBDE, would serveto better characterize the immunotoxic potential of PBDEs. PBDEs 3293. HEALTH EFFECTSNeurotoxicity. While a large body of evidence indicates that the developing nervous system is a targetfor PBDE toxicity, a limited amount of information is available on neurological effects of PBDEs in adultanimals. In a comprehensive neurotoxicity screen, no adverse effects were observed in rats exposed onceto pentaBDE at doses up to 1.2 mg/kg/day (Belles et al. 2010). In a neurobehavioral study, no exposure-related changes were observed in open-field behavior, anxiety-like behavior, or learning and memory in male rats exposed to low doses of pentaBDE at doses up to 0.015 mg/kg/day for 90 days (Daubie et al. 2011). In a higher dose study, impaired attention and inhibitory control were observed in male miceexposed to pentaBDE at 26.2 mg/kg/day for 125 days; no exposure-related changes were observed at17.5 mg/kg/day (Driscoll et al. 2009). Impaired learning and memory were observed in male rats exposed to tetraBDE at doses ≥0.1 for 30 days (Yan et al. 2012). For decaBDE, no studies evaluating neurological end points in adult rats or mice following acuteexposure were identified. Decreased anxiety behavior was observed in male mice exposed to decaBDEfor 15 days; however, no exposure-related changes were observed in a functional observation battery orlearning and memory (Heredia et al. 2012). In another study, no changes were observed in open-fieldbehavior of male rats exposed to decaBDE at doses up to 50 mg/kg/day 90 days (Wang et al. 2011b). No overt signs of neurotoxicity were observed in rats and mice exposed to decaBDE in estimated dietary doses as high as 16,000–19,000 mg/kg/day for 14 days, 8,000–9,000 mg/kg/day for 13 weeks, or 2,550– 7,780 mg/kg/day for 103 weeks (NTP 1986). Although the high doses and extended exposure durationsprovided opportunities for the induction and/or development of clinical signs, the study is limited by lackof testing for subtle behavioral changes and neurodevelopmental effects. Additional comprehensiveneurotoxicity batteries in adult animals exposed orally to repeat doses of PBDEs would better characterize the potential for PBDEs to cause neurotoxic effects in adults.Epidemiological and Human Dosimetry Studies. There are numerous epidemiological studiesevaluating potential associations between tissue PBDE concentrations and adverse health effects;however, none of the studies provided quantitative exposure information. Therefore, the availableepidemiological studies are not useful for quantitative risk assessment. Available studies indicate thatPBDE exposure may lead to neurodevelopmental effects (Adgent et al. 2014; Chao et al. 2011; Chen et al. 2014; Chevrier et al. 2016; Cowell et al. 2015; Ding et al. 2015; Eskenazi et al. 2013; Gascon et al. 2012;Herbstman et al. 2010; Hoffman et al. 2012; Roze et al. 2009; Sagiv et al. 2015; Shy et al. 2011; Vuong etal. 2016a). Evidence for associations between PBDE exposure and other effects in humans are inconsistent between studies, including observations of altered thyroid hormone levels (Abdelouahab etal. 2011, 2013; Bloom et al. 2008, 2014; Chevrier et al. 2010; Dallaire et al. 2009; Eggesbo et al. 2011; PBDEs 3303. HEALTH EFFECTSEguchi et al. 2015; Hagmar et al. 2001; Herbstman et al. 2008; Huang et al. 2014; Julander et al. 2005;Kicinski et al. 2012; Kim et al. 2011a, 2011d, 2012a, 2012b, 2013a; Leijs et al. 2012; Lignell et al. 2016; Makey et al. 2016; Mazdai et al. 2003; Shy et al. 2012; Stapleton et al. 2011; Turyk et al. 2008; Yuan etal. 2008; Wan et al. 2010; Wang et al. 2010; Xu et al. 2014a, 2014b, 2015a), male reproductive effects(Abdelouahab et al. 2011; Akutsu et al. 2008; Hagmar et al. 2001; Johnson et al. 2013; Meeker et al. 2009; Mumford et al. 2015; Turyk et al. 2008), female reproductive effects (Buck Louis et al. 2013; Chao et al. 2007, 2010; Chen et al.2011; Harley et al. 2010; Johnson et al. 2012; Karmaus et al. 2011; Wainman et al. 2016), non-neurological developmental effects (Agay-Shay et al. 2015; Carmichael et al. 2010;Chao et al. 2007; Chen et al. 2015; Erkin-Cakmak et al. 2015; Foster et al. 2011; Harley et al. 2011; Kim et al. 2011d, 2012a, 2012b, 2015; Koskenniemi et al. 2015; Leijs et al. 2008; Lopez-Espinosa et al. 2015;Ma et al. 2012a; Main et al. 2007; Mazdai et al. 2003; Meijer et al. 2012; Miranda et al. 2015; Müller etal. 2016; Ochiai et al. 2014; Peltier et al. 2015; Ren et al. 2011; Robledo et al. 2015a; Serme-Gbedo et al. 2016; Tan et al. 2009; Vuong et al. 2016b; Warembourg et al. 2016; Windham et al. 2015a; Wu et al.2010; Xu et al. 2015b), and risk for diabetes (Airaksinen et al. 2011; Lee et al. 2011; Lim et al. 2008; Smarr et al. 2016; Turyk et al. 2015). In other epidemiological studies, serum PBDE levels were notsignificantly associated with carotid atherosclerosis, risk of stroke, bone mineral density, immunefunction, non-Hodgkin’s lymphoma (Fernlof et al. 1997; Hardell et al. 1998; Kumar et al. 2014a, 2014b;Lee et al. 2012; Lind et al. 2012; Lindstrom et al. 1998;Weiss et al. 2006), breast cancer (Holmes et al.2014; Hurley et al. 2011), thyroid cancer (Aschebrook-Kilfoy et al. 2015), prostate cancer (Pi et al. 2016), reproductive effects in men (Toft et al. 2014), Polycystic Ovary Syndrome (Vagi et al. 2014), or uterinefibroids (Trabert et al. 2015). Epidemiological studies with quantitative estimates of exposure would be useful for quantitative riskassessment. Considering the possibility that PBDEs can be transferred to the fetus across the placenta andthat greater amounts might be transferred to nursing infants via breast milk, as well as evidence thatperinatal exposure to PCBs and other similar chemicals may induce subtle neurological damage andimmunological and thyroid effects in children, transgenerational studies would be particularlyinformative. Additional studies evaluating the potential link between pre- and peri-natal PBDE exposure and childhood behavior disorders, such as Autism Spectrum Disorder (ASD) and ADHD, have been specifically requested by the scientific community based upon potential links between these disorders andendocrine disruption (de Cock et al. 2012; Messer 2010). However, limitations that are likely to constrainepidemiological investigations, such as unmeasured PBDE exposure concentrations and lack of controls for confounding co-exposures, may be difficult to address. Studies that assess PBDE concentrations inserum or breast milk along with concentrations of other persistent organic pollutants (POPs), such as PBDEs 3313. HEALTH EFFECTSPCBs, PCDDs, and /or PCDFs, would be useful to evaluate responses that may be co-dependent on otherpersistent lipophilic agents.Biomarkers of Exposure and Effect.Exposure. PBDEs accumulate in adipose tissue, serum, and breast milk of the general population due to their lipophilic characteristics. Concentrations of PBDEs in breast milk are useful, non-invasive markers of maternal body burdens and of in utero and lactational exposures, but body burden assessments are limited by a lack of time-trend data for PBDEs in the milk of U.S. populations (Hooper and McDonald 2000). Breast milk monitoring programs would provide time-trend data that would verify whetherregulatory action to limit the use of PBDEs is reversing the previous trend of an exponential increase in PBDE concentrations in breast milk (Norén and Meironyté 1998, 2000). Studies on the predictive valueof concentrations of PBDEs in serum and adipose tissue could provide useful information for detection and monitoring of exposure. It should be noted, however, that solubilities in adipose and breast milk arelikely to vary with the congener. For example, decaBDE is much less soluble in adipose than pentaBDE.These differences must be considered when designing studies evaluating PBDE exposure. Recent studies have indicated that PBDE concentrations in hair may also be useful to estimate exposure levels (Aleksa etal. 2012a, 2012b; Malarvannan et al. 2013), although Zheng et al. (2011) reported that PBDE concentrations in hair were not correlated with PBDE concentrations in indoor dust from urban, e-waste,and rural areas in South China. Additional studies would be useful to validate this approach.A potential biomarker of exposure to PBDEs relates to their effect on the thyroid gland. Thyroid changesin rats and mice include reduced serum T4 levels, with no changes in serum TSH (Darnerud and Sinjari1996; Fowles et al. 1994; Hallgren and Darnerud 1998; WIL Research Laboratories 1984; Zhou et al. 2001, 2002). However, using thyroid changes as a biomarker may not be reliable, as thyroid changes arenot specific to exposure to PBDEs and the effects associated with the thyroid in non-clinical studies arelikely specific to the rodent and may or may not be directly relevant to the human. Additional studiescould characterize thyroid effects of PBDEs in humans and develop specific correlations between levels and duration of exposure and alterations in serum levels of T4.Effect. Biomarkers that could be used to characterize health effects caused specifically by exposure to PBDEs have not been identified. Additional information on the mechanisms of toxicity may suggest auseful biomarker of effect; however, at this time, there is little to suggest that such biomarkers exist. PBDEs 3323. HEALTH EFFECTSAbsorption, Distribution, Metabolism, and Excretion. No information was located regardingthe toxicokinetics of PBDEs in humans following controlled oral exposure, although information oftoxicokinetics has been gathered based on levels of PBDEs and metabolites in tissues of environmentallyexposed individuals and numerous animal studies and in vitro studies. Absorption studies in animals indicate that decaBDE is absorbed to a lesser degree than lower-brominated PBDEs. The most recent and best available estimates of oral absorption efficiencies indicate a range of70–85% for the tetraBDE (BDE 47), pentaDE (BDE 99, BDE 100), and hexaBDE (BDE 153, BDE 154), and 10–26% for decaBDE (BDE 209) (Chen et al. 2006; Hakk et al. 2002a, 2002b, 2009; Klasson Wehleret al. 2001; Morck and Klasson Wehler 2001; Morck et al. 2003; Örn and Klasson-Wehler 1998; Riu et al. 2008; Sandholm et al. 2003; Sanders et al. 2006a, 2006b; Staskal et al. 2005). Quantitative absorption studies in humans could corroborate the conclusions on oral uptake in animals that are based onelimination and excretion data.Distribution studies in animals indicate that lower-brominated BDE congeners, following absorption and an initial wide distribution, are preferentially accumulated in adipose tissues (Chen et al. 2006; El Dareeret al. 1987; Hakk et al. 2002a, 2009; Morck and Klasson-Wehler 2001; Morck et al. 2003; Norris et al. 1975a; NTP 1986; Örn and Klasson-Wehler 1998; Riu et al. 2008; Sanders et al. 2006a, 2006b; Staskal etal. 2005, 2006a). In comparison, decaBDE is more readily distributed to highly perfused tissues and lessreadily distributed to adipose tissues. Evidence for the transfer of PBDEs from pregnant mothers to thedeveloping fetus and for the transfer of PBDEs from maternal blood to breast milk and then to nursinginfants comes from a number of studies of PBDE concentrations in maternal and cord serum samples and breast milk samples from groups of non-occupationally exposed women. In general, the tetra- and pentabrominated PBDEs have been the predominant congeners detected in maternal and cord serum samplesand breast milk samples, but some recent studies assaying for a wider range of PBDE congeners have found evidence for distribution of hepta-, octa-, or decaBDEs into cord serum and breast milk (Antignacet al. 2009, 2008; Chen et al. 2013; Daniels et al. 2010; Hites 2004; Kawashiro et al. 2008; Li et al. 2013a; Malarvannan et al. 2013; Mazdai et al. 2003; Meijer et al. 2008; Park et al. 2011; Qiu et al. 2009;Schecter et al. 2010, 2006; Vizcaino et al. 2011; Wan et al. 2010). Maternal transfer of both lowerbrominated PBDEs and decaBDE has also been shown in animal studies (Cai et al. 2011; Zhang et al. 2011). The available studies appear adequate to characterize distribution of PBDEs.Current evidence indicates that CYP2B6-mediated metabolism of BDE 47 and BDE 99 produces multiplehydroxylated metabolites via hydroxylation and ether bond cleavage, based on in vitro studies with PBDEs 3333. HEALTH EFFECTShuman liver microsomes or hepatocytes and human recombinant CYPs (Erratico et al. 2012, 2013; Feo etal. 2013). The major metabolites of BDE 47 and BDE 99 formed by human liver microsomes were notthe same as those identified using rat liver microsomes (Erratico et al. 2013, 2012, 2011). Differentclasses of CYP enzymes appear to be involved in in vitro rat liver metabolism of BDE 47 and BDE 99:CYP1A1, CYP2A2, and CYP3A1 for BDE 47 and CYP1A1, CYP2A2, CYP2B1, and CYP3A1 for BDE99 (Erratico et al. 2011). Production of hydroxylated metabolites of BDE 153 (Lupton et al. 2009) and BDE 209 (Stapleton et al. 2009) has not been demonstrated with human liver microsomes or hepatocytes.It is uncertain if these latter findings are reflective of a limited in vivo capacity of humans to metabolize these BDE congeners or because the proper in vitro conditions for metabolizing these congeners were notprovided. Currently, studies of metabolism of BDE 47 and BDE 99 using in vitro human and rat systems have found evidence of metabolic oxidative debromination only with BDE 47 in human livermicrosomes. In contrast, a number of in vivo studies have found evidence for oxidative debromination by analysis of feces collected from rats exposed to BDE 47 (Marsh et al. 2006), BDE 99 (Hakk et al. 2002a), BDE 100 (Hakk et al. 2006), BDE 154 (Hakk et al. 2009), and BDE 209 (Morck et al. 2003; Sandholm etal. 2003). Additional metabolism studies would help to characterize the enzymes involved as well as the transformation of some congeners to biologically active hydroxylated BDEs. There are still data gaps in the toxicokinetics of decaBDE, including an incomplete understanding of the debromination of decaBDEto lower-brominated BDEs.The detection of PBDEs in human breast milk samples suggest that breast milk represents an eliminationroute of absorbed PBDEs in women (see Jakobsson et al. 2012 and Frederiksen et al. 2009 for reviews ofPBDE levels in breast milk); however, studies do not provide a clear account of the degree to whichPBDEs are cleared from the body during breastfeeding (Hooper et al. 2007; Jakobsson et al. 2012;LaKind et al. 2009; Thomsen et al. 2010). Rat studies indicate that ingested PBDEs are principallyexcreted in the feces with <2% of administered radioactivity excreted in the urine within 3 days of doseadministration (Chen et al. 2006; Hakk et al. 2002a, 2006; Morck et al. 2003; Norris et al. 1973, 1975b; Orn and Klasson-Wehler 1998; Sanders et al. 2006a; Riu et al. 2008). However, a different elimination pattern has been observed in mice, especially with BDE 47. In mice given single oral doses of14C-labeled PBDE congeners, fecal and urinary elimination were principal routes of elimination forBDE 47 (Orn and Klasson-Wehler 1998; Sanders et al. 2006a), whereas fecal elimination appeared to bemore important than urinary elimination with BDE 99 (Chen et al. 2006) and BDE 153 (Sanders et al.2006b). Complementary studies with female C57BL/6J given single intravenous 1-mg/kg doses of14C-labeled BDE 47, BDE 99, BDE 100, or BDE 153 also indicate that the importance of urinaryexcretion in mice is congener specific (Staskal et al. 2006b). Quantitative studies in humans could PBDEs 3343. HEALTH EFFECTSdetermine which excretion route is more relevant for humans (feces or urine) for different congenersbased on elimination and excretion data. Additionally, further studies regarding the degree of elimination in breast milk during breast feeding could better characterize exposure risks for breastfeeding infants.Comparative Toxicokinetics. Insufficient data are available to determine whether there are qualitative differences in the toxicokinetic disposition of PBDEs between humans and animals and amonganimal species. However, elimination studies in rats and mice (discussed above) highlight thattoxicokinetics may differ between species. Differences are likely to be dependent on the specific congener or mixture studied, and pharmacokinetic modeling studies could help to determine the validityof extrapolating data. Most of the available toxicokinetic studies of PBDEs have been performed in rats,and studies in other species could help to ascertain the most relevant animal model.Methods for Reducing Toxic Effects. The mechanism by which PBDEs enter the blood stream is not known, there are no established methods for reducing body burden of PBDEs, and the mechanisms oftoxic action of PBDEs are incompletely understood. A more complete characterization of the cytosolicAhR protein and understanding of physiological effects of receptor blockage would be useful for thepossible identification of blockers of AhR-mediated toxic effects. Further studies aimed at elucidating thenonreceptor-mediated mechanism of action of some PBDEs would also be valuable.Children’s Susceptibility. Data needs relating to both prenatal and childhood exposures, anddevelopmental effects expressed either prenatally or during childhood, are discussed in detail in theDevelopmental Toxicity subsection above.Body burden data, as well as intake modeling, suggest that infants and toddlers have higher exposures toPBDEs as compared to older children or adults (EPA 2010; Lorber 2008; Trudel et al. 2011; Wong et al. 2013). Several epidemiological studies have reported results suggestive of an effect of PBDE onneurodevelopment in children (Adgent et al. 2014; Chao et al. 2011; Chen et al. 2014; Chevrier et al.2016; Cowell et al. 2015; Ding et al. 2015; Eskenazi et al. 2013; Gascon et al. 2012; Herbstman et al.2010; Hoffman et al. 2012; Roze et al. 2009; Sagiv et al. 2015; Shy et al. 2011; Vuong et al. 2016a), and these findings are supported by developmental studies in animals (Biesemeier et al. 2011; Blanco et al. 2013; Branchi et al. 2001, 2002, 2005; Cheng et al. 2009; Eriksson et al. 2001, 2002, 2006; Fischer et al. 2008; Fujimoto et al. 2011; Gee and Moser 2008; He et al. 2009, 2011; Johansson et al. 2008; Koenig etal. 2012; Kuriyama et al. 2004, 2005; Rice et al. 2007; Sand et al. 2004; Ta et al. 2011; Viberg et al. 2002,2003a, 2003b, 2004a, 2004b, 2005, 2006, 2007; Woods et al. 2012; Xing et al. 2009). Epidemiological PBDEs 3353. HEALTH EFFECTSstudies in infants are inconclusive regarding thyroid effects and exposure to PBDEs (Abdelouahab et al. 2013; Herbstman et al. 2010; Kim et al. 2011a, 2012a; Lin et al. 2011; Mazdai et al. 2003; Shy et al. 2012); however, numerous animal studies indicated that developmental exposure to lower-brominatedPBDEs result in thyroid hormone alterations (Blanco et al. 2013; Bondy et al. 2011, 2013; Ellis-Hutchings et al. 2006; Kodavanti et al. 2010; Kuriyama et al. 2007; Miller et al. 2012; Poon et al. 2011;Shah et al. 2011; Skarman et al. 2005; Szabo et al. 2009; Wang et al. 2011a; Zhang et al. 2009; Zhou et al.2002). Additional studies would better characterize the potential susceptibility of children to the effects of PBDEs on the thyroid and neurodevelopment, particularly considering the possibility that these effectsare related to the dependence of central nervous system development on thyroid hormones.Data from two human epidemiological studies suggest that PBDE exposure may alter reproductive systemdevelopment in boys (Main et al. 2007; Meijer et al. 2012). In contrast, other studies found no associations between serum or adipose tissue PBDE concentrations and reproductive development(Carmichael et al. 2010; Koskenniemi et al. 2015; Leijs et al. 2008). These findings, along with limited evidence of reproductive effects in animals exposed to lower-brominated PBDEs during development(Kuriyama et al. 2005; Talsness et al. 2005, 2008), indicate that additional studies of reproductivedevelopment in young animals would help to more fully assess children’s susceptibility to PBDEs.No information is available regarding the immunosuppressive potential of PBDEs in children, but serumlevels of BDE 28 and 209 were found associated with an increased risk of asthma in a study of 3–6-yearold Chinese children (Meng et al. 2016). Limited evidence in young animals exposed to PBDEs (Bondyet al. 2013; Watanabe et al. 2008, 2010b) indicates that additional studies of immune competence in developing animals would also help to more fully assess children’s susceptibility to PBDEs.A recent small study of Swedish toddlers (n=22) found good correlations between concentrations of somePBDEs in serum and in feces, suggesting that fecal analysis, a noninvasive test, might be a good alternative for biomonitoring PBDEs in toddlers (Sahlström et al. 2015). Larger studies would bevaluable to improve the predictive power of the statistical analyses performed in this study. In addition, itwould be helpful to determine whether the method is applicable to infants and children of different ages.Child health data needs relating to exposure are discussed in Section 6.8.1, Identification of Data Needs:Exposures of Children. PBDEs 3363. HEALTH EFFECTS3.12.3 Ongoing Studies Fifty-six ongoing research efforts have been identified that may provide data related to the toxic actionsof PBDEs in humans (RePORTER 2014, 2016). These projects are summarized in Table 3-7.Animal Studies. Dr. Charles Herbert of the Southern Research Institute is conducting chronic studies ofthe commercial pentaBDE mixture DE-71 in rats and mice under contract with NTP and Dr. Lu Taylor ofthe Battelle Memorial Institute is conducting a perinatal study BDE 47 in CB6F1-Tg(HRAS2) transgenic mice under contract with NTP.Several animal studies are investigating neurobehavioral changes in animals exposed to PBDEs. Fourongoing studies are investigating behavioral and biochemical end points relevant to ADHD (Drs. Susan Schantz and Paula Eubig, University of Illinois Urbana-Champaign; Dr. Richard Seegal, Wadsworth Center). To assess the role of the immune system and its interaction with environmental contaminants inautism and other neurodevelopmental disorders, Dr. Isaac Pessah of the University of California at Davis is investigating the effects of perinatal PBDE exposure on brain development, complex social behaviors,and immune system function in mouse strains with low (C57BL/6J) or high (SJL mice) susceptibility to autoimmunity. Using an established mouse model, Dr. Margarita Behrens, from the Salk Institute forBiological Sciences, will examine the effects of exposure to PBDE on autism spectrum disorder. Dr.Helen Sable of the University of Memphis is examining if developmental PBDE exposure in rats alterdopamine receptor expression and enhance behavioral sensitization following psychostimulant exposure(as seen with PCBs).Dr. Deena Small of the University of New England is examining bone growth and remodeling in mice exposed to pentaBDE from PND 1 to 60. In addition, Dr. Small will use cell culture-based assays thatmeasure gene expression, enzyme activity and calcium deposition in cultured bone cell lines exposed to the pentaBDEs.Toxicokinetic Studies. Dr. Tracey Woodruff of the University of California, San Francisco is measuringconcentrations of PBDEs and OH-PBDEs in human maternal and fetal biological specimens from women undergoing voluntary, second trimester pregnancy terminations and generating original human data on whether fetal exposures to PBDEs alter gene expression of cytochrome P450 (CYP) enzymes in thesecond-trimester human fetal liver and placenta. Dr. James Olson of SUNY Buffalo is conducting a qualitative and quantitative characterization of the human CYP-specific in vitro metabolism of PBDEs 3373. HEALTH EFFECTSTable 3-7. Ongoing Research for PBDEs in HumansInvestigator Affiliation Research description SponsorBaccarelli, A Harvard School of Public Prenatal exposure to PBDEs and visual skills, NIEHSHealth attention, and fine motor skills in childrenBarr, DB Emory University Prenatal exposure to PBDEs and birth NIHESoutcomesBlumberg, B University of California- Endocrine disrupter modulation of the steroid NIEHSIrvine and xenobiotic receptor, SXR, in developmentand lymphomagenesisCaudle, WM Emory University Vesicular monoamine transporter 2 as a NIEHSmediator of PBDE neurotoxicityChen, A University of Cincinnati Longitudinal study of exposure to PBDEs and NIEHSchild neurobehaviorChen, S Beckman Research Determine role and mechanism of PBDEs on NIEHSInstitute/City of Hope development of breast cancer during the menopausal transitionCosta, LG University of Washington Low-level exposure to PBDEs: Testing the NIEHShermetic and epigenetic hypothesesCroen, LA Kaiser Foundation Research Prenatal and neonatal biologic markers for NIEHSInstitute autismDarrow, L Emory University PBDE body burdens, house dust NIEHSconcentrations, and associations with thyroid hormoneDi Giulio, R Duke University Thyroid metabolism disruption key in toxicant- NIEHSinduced developmentEskenazi, B University of California Pesticides and PBDEs on neurobehavior NIEHSBerkeleyEskenazi, B University of California PBDEs, DDT, and neurodevelopment in NIEHSBerkeley school-aged Mexican-American childrenEskenazi, B University of California DDT and PBDE exposure, puberty onset, and NIEHSBerkeley neurodevelopment in Mexican-American girlsEubig, P University of Illinois Urbana- Effects of PCBs and PBDEs on three distinct NIEHSChampaign components of response inhibitionFerguson, PL University of South Carolina Mechanisms of xenoestrogen stress: a NIEHSat Columbia proteomic and functional genomic approachFitzgerald, EF State University of New York Persistent organic pollutants and cognitive NIEHSat Albany decline in the elderlyGetahun, D Kaiser Foundation Research Flame retardant and adverse perinatal NIEHSInstitute outcomeGiese, RW Northeastern University Discovery of xenobiotics associated with NIEHSpreterm birthHauser, RB Harvard School of Public Maternal and paternal flame retardant NIEHSHealth exposure, impact on fertility and pregnancyHerbert, C Southern Research Institute Studies to evaluate the toxicologic and NIEHScarcinogenic potentialHerbstman, JB Columbia University Pre- and postnatal PBDE exposure, thyroid NIEHShormones, and neurodevelopment PBDEs 3383. HEALTH EFFECTSTable 3-7. Ongoing Research for PBDEs in HumansInvestigator Affiliation Research description SponsorEpidemiology and the environment in autism NIEHSThe CHARGE study: Childhood Autism Riskfrom Genetics and EnvironmentNIEHSAutism risk, prenatal environmentalexposures, and pathophysiologic markersNIEHSEpigenetic effects of DDT/E and PBDEs onpubertyNIEHSEnvironmental determinants of autoimmunityamong African Americans in coastal South NIEHSCarolinaDevelopmental exposure to PBDEs and long- NIEHSlasting modifications of drug metabolism in childrenEpigenetic interaction of MECP2 and organic NIEHSpollutants in neurodevelopmentMethylomic and genomic impacts of organic NIEHSpollutants in Dup15q syndromeMechanisms of inflammation in gestational NIEHSmembranesLongitudinal investigation of fertility and the NICHDenvironmentProtecting the health of future generations: NIEHSassessing and preventing exposuresEffects of endocrine disrupting chemicals and NIEHSchronic psychosocial stress on fetal growthBioactivation of PBDEs by human cytochrome NIEHSP-450Calcium signaling defects in autism NIEHSNeurodevelopmental toxicology of autism NIEHSEnvironment chemicals and gynecologic NICHDhealth Exposure Assessment for childhood leukemia NIEHSPolybrominated diphenyl ether effects on NIEHShuman neuronal developmentAssessment of psychostimulant addiction risk NIEHSfollowing developmental PCB exposure Mechanisms of xenoestrogen stress: a NIEHSproteomic and functional genomic approachHertz-Picciotto,IHertz-Picciotto,IHertz-Picciotto,IHolland, NTKamen, DLKlaassen, CDLasalle, JMLasalle, JMLoch-Caruso,RKLouis, GMiller, PKMorello-Frosch, RAOlson, JRPessah, INPessah, INPeterson, MUniversity of California atDavisUniversity of California atDavisUniversity of California atDavisUniversity of California BerkeleyMedical University of SouthCarolinaUniversity of WashingtonUniversity of California atDavisUniversity of California atDavisUniversity of MichiganEunice Kennedy ShriverNational Institute of Child Health & Human DevelopmentAlaska Community Actionon Toxics (ACAT)University of California, San FranciscoState University of New Yorkat BuffaloUniversity of California atDavisUniversity of California atDavisUniversity of UtahRappaport, SM University of California BerkeleyRobinson, JF University of California, San FranciscoSable, HJ University of MemphisSabo-Attwood, University of South Carolina TL at Columbia PBDEs 3393. HEALTH EFFECTSTable 3-7. Ongoing Research for PBDEs in HumansInvestigator Affiliation Research description SponsorSchantz, SLSchantz, SLSeegal, RFSmall, DJStapleton, HMStapleton, HMTaylor, LUTuryk, MEVan de Water,JAWapner, RWebster, TFWoodruff, TJWoodruff, TJZhang, YZota, ARUniversity of Illinois Urbana-ChampaignUniversity of Illinois Urbana-ChampaignWadsworth CenterUniversity of New EnglandDuke UniversityDuke UniversityBattelle Memorial InstituteUniversity of Illinois atChicagoUniversity of California atDavisColumbia University HealthSciencesBoston University MedicalCampusUniversity of California, San FranciscoUniversity of California, San FranciscoYale UniversityGeorge Washington UniversityPCBs, PBDEs, hearing loss, and attention/impulsivity: mechanistic studies inanimalsHealth effects of PCB exposure fromcontaminated fishDevelopmental neuroendocrine effects ofPCBs and PBDEsEffect of polybrominated diphenyl ether flame retardant exposure on osteogenesisChildren’s exposure to flame retardants:effects on thyroid hormone regulationDeiodinase activity as a biomarker ofresponse to brominated flame retardantsStudies to evaluate the toxicological potentialof test articlesDiabetes and persistent organic pollutantsImmune environment interaction and neurodevelopmentEndocrine disruption in pregnant women:thyroid disruption and infant developmentMeasuring human exposure to PBDEsHuman maternal/fetal exposures to PBDEsand their metabolites during developmentMid-gestational exposure to endocrine disrupting chemicals and effects on placentaldevelopmentPolyhalogenated aromatic hydrocarbons and thyroid cancer risk in the Department ofDefense Serum Repository (DoDSR) cohortRole of endocrine-disrupting chemicals and social stress on perinatal outcomesNIEHSATSDRNIEHSNIEHSNIEHSNIEHSNIEHSNIEHSNIEHSNIEHSNIEHSNIEHSNIEHSNIEHSNIEHSSource: RePORTER 2014, 2016NIEHS = National Institute of Environmental Health Sciences; PBDE = polybrominated diphenyl ether;PCB = polychlorinated biphenyl PBDEs 3403. HEALTH EFFECTS2,2',4,4'-tetraBDE (BDE 47), 2,2',4,5'-tetraBDE (BDE 49), 2,2',4,4',5-pentaBDE (BDE 99), and 2,2',4,4',6- pentaBDE (BDE 100) as well as quantify OH-PBDEs in human milk and serum and assess thepotential impact of CYP2B6 genotype on the body burden of PBDEs. Dr. Heather Stapleton of DukeUniversity is identifying the products of hepatic metabolism of PBDEs (no further details available).Mechanistic Studies. Several studies are investigating mechanisms of neurotoxicity. At the Universityof California at Davis, Dr. Janine Lasalle is investigating epigenetic changes in the genome of Mecp2mutant mouse models of Rett syndrome and autism following in vivo exposure to BDE 47, Drs. JanineLasalle and Isaac Pessah are investigating potential mechanisms of PBDE toxicity in susceptible neuronalcell models (chromosome 15q11-13 duplication syndrome [Dup15q] or fragile X syndrome [FMRI]), and Dr. Judy Van de Water is investigating mechanisms of PBDE toxicity in peripheral blood mononuclearcells from children with autism. Dr. Joshua Robinson of the University of California San Francisco isinvestigating the effects in vitro exposure to BDE 47 and BDE 99 on neuronal differentiation and geneexpression in human embryonic stem cells. Dr. Lucio Costa of University of Washington is investigatinggenetic and epigenetic changes in cultured mouse neurons exposed to low, environmentally-relevantconcentrations of BDE 47. Dr. William Caudle of Emory University is investigating the potential mechanisms by which the commercial pentaBDE mixture DE-71 disrupts VMAT2, which is a keymediator of cytosolic dopamine.Additional studies are investigating mechanisms of thyroid hormone disruption. Dr. Heather Stapleton ofDuke University is investigating the effects of PBDEs and their metabolites on intra- and extra-cellularthyroid hormone regulation in vitro to elucidate the mechanisms of action for thyroid toxicity. Dr. DeenaSmall of the University of New England is measuring thyroid hormone receptor- mediated transcription and thyroid hormone receptor binding of pentaBDEs in vitro. Dr. Richard Di Guilio of Duke Universityis investigating the potential role of inhibition of diodinases by PBDEs in altered thyroid hormonehomeostasis, using zebrafish as a model.Other mechanistic studies are investigating mechanisms of xenoestrogen stress (Drs. P. Lee Ferguson andTara Sabo-Attwood of the University of South Carolina at Columbia), potential mechanisms behind the etiology of inflammation of extra-placental gestational membranes and associated release of cytokines and prostaglandins (which are associated with preterm birth) (Dr. Rita Loch-Caruso of the University ofMichigan), the potential role of the steroid and xenobiotic receptor SXR in the development of lymphomaand leukemia in individuals exposed to PBDEs (Dr. Bruce Blumberg of the University of California-Irvine), and development of biomarkers of mitochondrial function in primary mouse hepatocytes exposed PBDEs 3413. HEALTH EFFECTSto BDE 47 using metabolomic analysis (Dr. Hockenberry from the Fred Hutchison Cancer ResearchCenter. PBDEs 3423. HEALTH EFFECTSThis page is intentionally blank. PBDEs 3434. CHEMICAL AND PHYSICAL INFORMATION4.1 CHEMICAL IDENTITYPBDEs are a class of structurally similar brominated hydrocarbons, in which 2–10 bromine atoms are attached to the diphenyl ether molecule. Monobrominated structures (i.e., one bromine atom attached to the molecule) are often included when describing PBDEs. The general chemical structure of PBDEs is shown below:O BrnBrm 1' 2' 3' 4' 5' 6' 1 2 3 4 5 6 where m + n = 1 to 10It can be seen from the structure that a large number of brominated compounds are possible. The209 possible compounds for PBDEs are called “congeners”. However, the number of PBDE congenersthat actually exist in commercial PBDE mixtures are much less compared to PCBs. Typically, only asubset of the 209 possible congeners is observed for PBDEs. PBDEs can also be categorized by degree ofbromination. The term “homolog” is used to refer to all PBDEs with the same number of bromines (e.g., tribromodiphenyl ether or triBDE refers to PBDEs containing only three bromine atoms). Based on thenumber of bromine substituents, there are 10 homologous groups of PBDEs (monobrominated through decabrominated). Each homologous group contains one or more congeners. The mono-, di-, tri-, tetra-, penta-, hexa-, hepta-, octa-, nona-, and decabromo-congeners can exist in 3, 12, 24, 42, 46, 42, 24, 12, 3, and 1 forms, respectively. Homologs with different substitution patterns are referred to as isomers. Forexample, the group of dibromodiphenyl ether or diBDE homologs contains 12 isomers. The numberingsystem for PBDEs is also shown above. The structures of representative PBDE molecules appear similarwhen drawn in one dimension. However, there are important three-dimensional differences in theirstructures due to the ether linkage and location/number of halogen atoms. The ortho positions of thearomatic rings must be nonhalogen-substituted for a diphenyl ether molecule to assume a planar or nearplanar configuration. Halogen substitution of the diphenyl ether molecule in the ortho position (2,2’,6,6’)will force the aromatic rings orthogonal to one another (e.g., the phenyl rings will be positioned in spacewith a dihedral angle >0°). This is particularly evident for decabromodiphenyl ether, which is predicted to have a dihedral angle of ~90° and a high barrier to rotation around the ether linkage preventing this PBDEs 3444. CHEMICAL AND PHYSICAL INFORMATIONmolecule from assuming a planar configuration. The benzene rings of non-ortho substituted PBDEs mayassume a small dihedral angle (in which the dihedral angle is small, but >0°) or “near” planarconfiguration. These molecules are referred to as planar or coplanar congeners (Hardy 2002a). Like PCBs, the 209 congeners for PBDEs are arranged in ascending numerical order using a numberingsystem developed by Ballschmiter and Zell (1980) that follow the IUPAC rules of substituentcharacterization in biphenyls. The resulting numbers assigned by Ballschmiter and Zell (which are alsoreferred to as congener, IUPAC, or BZ numbers) are widely used for identifying individual congeners ofPBDEs. For example, the PBDE congener, 2,2',4,4'-tetraBDE may be referred to as BDE 47 in thisdocument. The identities of several PBDE congeners are shown in Table 4-1 (WHO 1994a, 1994b).In the United States, Albemarle Corporation and Great Lakes Chemical Corporation previously marketedmixtures of PBDEs under trade names (e.g., DE-60F, DE-61, DE-62, DE-71, for pentaBDE mixtures;DE-79 for octaBDE mixtures; and DE 83R, Saytex 102E for decaBDE mixtures). The Great Lakes Corporation merged with Crompton Chemical Corporation and was renamed Chemtura, which produced decaBDE under the brand names AZUB DB-40, AZUB DB-65, AZUB 2DA-65, and AZUB 3DA-65 (EPA 2010). There were also several trade names used by producers from Europe and Japan for the BDEmixtures. The chemical identities of commercial mixtures of penta-, octa-, and decaBDEs are listed inTable 4-2 (WHO 1994a). La Guardia et al. (2006) published detailed congener composition profiles of penta-, octa-, and decaBDE flame retardant mixtures; 39 discrete PBDEs were found in the six commercial products evaluated by GC/MS electron ionization (EI) and electron-capture negative ionization (ECNI).Various synonyms and abbreviations for PBDEs exist in the literature and are shown below:polybrominated biphenyl ethers = polybromobiphenyl ethers = PBBEpolybrominated biphenyl oxides = polybromobiphenyl oxides = PBBEspolybrominated diphenyl ethers = polybromodiphenyl ethers = PBDEs or PBDPEspolybrominated diphenyl oxides = polybromodiphenyl oxides = PBDOs or PBDPOs PBDEs 3454. CHEMICAL AND PHYSICAL INFORMATIONTable 4-1. Chemical Identity of Polybrominated Diphenyl Ether (PBDE)CongenersaIUPAC Numberb Compound/substituents CAS numberc Biphenyl 92-52-4MonoBDE 101-55-31 22 33 4456789101112131415DiBDE 2050-47-72,2’2,32,3’2,42,4’2,52,63,3’3,43,4’3,54,4’TriBDE 49690-94-01617181920212223242526272829303132333435363738392,2’,32,2’,42,2’,52,2’,62,3,3’2,3,42,3,4’2,3,52,3,62,3’,42,3’,52,3’,62,4,4’2,4,52,4,62,4’,52,4’,62’,3,42’,3,53,3’,43,3’,53,4,4’3,4,53,4’,5 PBDEs 3464. CHEMICAL AND PHYSICAL INFORMATIONTable 4-1. Chemical Identity of Polybrominated Diphenyl Ether (PBDE)CongenersaIUPAC Numberb Compound/substituents CAS numberc 404142434445464748495051525354555657585960616263646566676869707172737475767778798081TetraBDE 40088-47-92,2’,3,3’2,2’,3,42,2’,3,4’2,2’,3,52,2’,4,5’2,2’,3,62,2’,3,6’2,2’,4,4'2,2’,4,52,2’,4,5'2,2’,4,62,2’,4,6'2,2’,5,5'2,2’,5,6'2,2’,6,6'2,3,3’,42,3,3’,4’2,3,3’,52,3,3’,5’2,3,3’,62,3,4,4’2,3,4,52,3,4,62,3,4’,52,3,4’,62,3,5,62,3’,4,4’2,3’,4,52,3’,4,5’2,3’,4,62,3’,4’,52,3’,4’,62,3’,5,5’2,3’,5’,62,4,4’,52,4,4’,62’,3,4,53,3’,4,4’3,3’,4,53,3’,4,5’3,3’,5,5’3,4,4’,5 PBDEs 3474. CHEMICAL AND PHYSICAL INFORMATIONTable 4-1. Chemical Identity of Polybrominated Diphenyl Ether (PBDE)CongenersaIUPAC Numberb Compound/substituents CAS numberc 828384858687888990919293949596979899100101102103104105106107108109110111112113114115116117118119120121122123124125126127PentaBDE 32534-81-92,2’,3,3’,42,2’,3,3’,52,2’,3,3’,62,2’,3,4,4’2,2’,3,4,52,2’,3,4,5’2,2’,3,4,62,2’,3,4,6’2,2’,3,4’,52,2’,3,4’,62,2’,3,5,5’2,2’,3,5,62,2’,3,5,6’2,2’,3,5’,62,2’,3,6,6’2,2’,3’,4,52,2’,3’,4,62,2’,4,4’,52,2’,4,4’,62,2’,4,5,5’2,2’,4,5,6’2,2’,4,5’,62,2’,4,6,6’2,3,3’,4,4’2,3,3’,4,52,3,3’,4’,52,3,3’,4,5’2,3,3’,4,62,3,3’,4’,62,3,3’,5,5’2,3,3’,5,62,3,3’,5’,62,3,4,4’,52,3,4,4’,62,3,4,5,62,3,4’,5,62,3’,4,4’,52,3’,4,4’,62,3’,4,5,5’2,3’,4,5’,62’,3,3’,4,52’,3,4,4’,52’,3,4,5,5’2’,3,4,5,6’3,3’,4,4’,53,3’,4,5,5’ PBDEs 3484. CHEMICAL AND PHYSICAL INFORMATIONTable 4-1. Chemical Identity of Polybrominated Diphenyl Ether (PBDE)CongenersaIUPAC Numberb Compound/substituents CAS numberc 128129130131132133134135136137138139140141142143144145146147148149150151152153154155156157158159160161162163164165166167168169HexaBDE 36483-60-02,2’,3,3’,4,4’2,2’3,3’,4,52,2’,3,3’,4,5’2,2’,3,3’,4,62,2’,3,3’,4,6’2,2’,3,3’,5,5’2,2’,3,3’,5,62,2’,3,3’,5,6’2,2’,3,3’,6,6’2,2’,3,4,4’,52,2’,3,4,4’,5’2,2’,3,4,4’,62,2’,3,4,4’,6’2,2’,3,4,5,5’2,2’,3,4,5,62,2’,3,4,5,6’2,2’,3,4,5’,62,2’,3,4,6,6’2,2’,3,4’,5,5’2,2’,3,4’,5,62,2’,3,4’,5,6’2,2’,3,4’,5’,62,2’,3,4’,5,6’2,2’,3,5,5’,62,2’,3,5,6,6’2,2’,4,4’,5,5’2,2’,4,4’,5,6’2,2’,4,4’,6,6’2,3,3’,4,4’,52,3,3’,4,4’,5’2,3,3’,4,4’,62,3,3’,4,5,5’2,3,3’,4,5,62,3,3’,4,5’,62,3,3’,4’,5,5’2,3,3’,4’,5,62,3,3’,4’,5’,62,3,3’,5,5’,62,3,4,4’,5,62,3’,4,4’,5,5’2,3’,4,4’,5’,63,3’,4,4’,5,5’ PBDEs 3494. CHEMICAL AND PHYSICAL INFORMATIONTable 4-1. Chemical Identity of Polybrominated Diphenyl Ether (PBDE)CongenersaIUPAC Numberb Compound/substituents CAS numberc 170171172173174175176177178179180181182183184185186187188189190191192193HeptaBDE 68928-80-32,2’,3,3’,4,4’,52,2’,3,3’,4,4’,62,2’,3,3’,4,5,5’2,2’,3,3’,4,5,62,2’,3,3’,4,5,6’2,2’,3,3’,4,5’,62,2’,3,3’,4,6,6’2,2’,3,3’,4’,5,62,2’,3,3’,5,5’,62,2’,3,3’,5,6,6’2,2’,3,4,4’,5,5’2,2’,3,4,4’,5,62,2’,3,4,4’,5,6’2,2’,3,4,4’,5’,62,2’,3,4,4’,6,6’2,2’,3,4,5,5’,62,2’,3,4,5,6,6’2,2’,3,4’,5,5’,62,2’,3,4’,5,6,6’2,3,3’,4,4’,5,5’2,3,3’,4,4’,5,62,3,3’,4,4’,5’,62,3,3’,4,5,5’,62,3,3’,4’,5,5’,6194195196197198199200201202203204205OctaBDE2,2’,3,3’,4,4’,5,5’2,2’,3,3’,4,4’,5,62,2’,3,3’,4,4’,5’,62,2’,3,3’,4,4’,6,6’2,2’,3,3’,4,5,5’,62,2’,3,3’,4,5,6,6’2,2’,3,3’,4,5,6,6’2,2’,3,3’,4,5’,6,6’2,2’,3,3’,5,5’,6,6’2,2’,3,4,4’,5,5’,62,2’,3,4,4’,5,6,6’2,3,3’,4,4’,5,5’,632536-52-0NonaBDE 63936-56-12062072082,2’,3,3’,4,4’,5,5’,62,2’,3,3’,4,4’,5,6,6’2,2’,3,3’,4,5,5’,6,6’ PBDEs 3504. CHEMICAL AND PHYSICAL INFORMATIONTable 4-1. Chemical Identity of Polybrominated Diphenyl Ether (PBDE)CongenersaIUPAC Numberb Compound/substituents CAS numberc DecaBDE 1163-19-5209 2,2’,3,3’,4,4’,5,5’,6,6’aWHO 1994a bBallschmiter and Zell 1980cNo CAS numbers were identified for the individual PBDE congeners.BDE = brominated diphenyl ether; CAS = Chemical Abstracts Service; IUPAC = International Union of Pure and Applied Chemistry PBDEs 3514. CHEMICAL AND PHYSICAL INFORMATIONTable 4-2. Chemical Identity of Technical Polybrominated Diphenyl Ethers(PBDEs)Characteristic Pentabromodiphenyl ether Octabromodiphenyl ether Decabromodiphenyl etherSynonym(s) Pentabromodiphenyl ether;pentabromodiphenyl oxide;pentabromobiphenyl oxide;benzene, 1,1-oxybis,pentabromo derivativeOctabromodiphenyl ether;Octabromodiphenyl oxide;octabromobiphenyl oxide;benzene, octabromo derivative; phenyl ether,octabromo derivativeDecabromodiphenyl ether;decabromodiphenyl oxide;decabromobiphenyl oxide;benzene, 1,1’-oxybis(2,3,5,6,-penta-bromo-) ether, bis(pentabromophenyl);Registered trade nameDE 71; Bromkal 70-5 DE;FR 1205/1215; Bromkal 70;Bromkal G1;Pentabromprop; Tardex 50;Tardex 50 L; Saytex 115Bromkal 79-8DE; DE 79;FR 143; Tardex 80;FR 1208; Adine 404;Saytex 111FR-300 BA; DE-83-RTM;Saytex 102; Saytex 102E;FR-1210; Adine 505;AFR 1021; Berkflam B10E;BR55N; Bromkal 81;Bromkal 82-0DE; Bromkal 83-10 DE; CalibanF/R-P 39P; Caliban F/R-P 44; Chemflam 011; DE 83;DP 10F; EB 10FP;EBR 700; Flame CutBR 100; FR P-39; BR 100;FR 330BA; FR P-39;FRP 53; FR-PE; FR-PE(H);Planelon DB 100;Tardex 100; NC-1085;HFO-102; Hexcel PF1;Phoscon Br-250Chemical formula C12H10-yBryO where y=4–6 C12H10-yBryO where y=6–9 C12Br10OChemical structureIdentification numbers:CAS registry 32534-81-9 32536-52-0 1163-19-5NIOSH RTECS No data No data No dataEPA hazardous No data No data No datawasteOHM/TADS No data No data No dataDOT/UN/IMCOshippingNo data No data No dataHSDB 7109 7110 2911NCI No data No data No dataSource: WHO 1994aCAS = Chemical Abstracts Services; DOT/UN/NA/IMCO = Department of Transportation/United Nations/North America/International Maritime Dangerous Goods Code; EPA = Environmental Protection Agency;HSDB = Hazardous Substances Data Bank; NCI = National Cancer Institute; NIOSH = National Institute forOccupational Safety and Health; OHM/TADS = Oil and Hazardous Materials/Technical Assistance Data System;RTECS = Registry of Toxic Effects of Chemical Substances PBDEs 3524. CHEMICAL AND PHYSICAL INFORMATIONFor consistency in this document, polybrominated diphenyl ethers or PBDEs will be used to identify thisclass of chemicals. The PBDE homologs are abbreviated as follows in this document:dibromodiphenyl ether = DiBDE = diBDEtribromodiphenyl ether = TrBDE = triBDEtetrabromodiphenyl ether = TeBDE = tetraBDEpentabromodiphenyl ether = PeBDE = pentaBDEhexabromodiphenyl ether = HxBDE = hexaBDEheptabromodiphenyl ether = HpBDE = heptaBDEoctabromodiphenyl ether = OcBDE = octaBDEnonabromodiphenyl ether = NoBDE = nonaBDEdecabromodiphenyl ether = DeBDE = decaBDE4.2 PHYSICAL AND CHEMICAL PROPERTIES Information found in the literature regarding the physical and chemical properties of selected technicalPBDE mixtures is presented in Table 4-3. Recent information regarding the vapor pressure, watersolubility, Henry’s Law constant, and log Kow of some PBDE congeners is presented in Table 4-4. Commercially available product mixtures of PBDEs (see Table 4-2) are not pure substances, but instead are mixtures of congeners. For example, the commercial mixture pentaBDE denotes the main componentof the mixture contains the pentaBDE homolog. However, the commercial pentaBDE mixture actuallycontains tetraBDE (24–38%) and pentaBDE (50–62%) homologs with small amounts of hexaBDE (4– 8%) and trace amounts of triBDE (0–1%) homologs. In this document, the commercial mixture ofpentabromodiphenyl ether may be called “the commercial pentaBDE mixture,” “technical pentaBDE,” or“technical PeBDE” to distinguish this mixture of homologs from the pentaBDE homolog, which refers to PBDEs with only five bromine atoms (see Section 4.1). Commercial octaBDE is a mixture of hexa-, hepta-, octa-, and nonaBDE homologs with trace amounts of decaBDE (i.e., BDE 209). In this document, the commercial mixture of octabromodiphenyl ether may be called “the commercial octaBDE mixture,” “technical octaBDE,” or “technical OBDE” to distinguish this mixture of different homologs from theoctaBDE homolog, which refers to PBDEs with only eight bromine atoms (see Section 4.1). Thecomposition of commercial decabromodiphenyl ether is 97% of the decaBDE (i.e., BDE 209); theremainder is nonaBDE homologs and trace amounts of octaBDE homologs (WHO 1994a). In thisdocument, commercial decabromodiphenyl ether may be called “the commercial decaBDE mixture,” PBDEs 3534. CHEMICAL AND PHYSICAL INFORMATIONTable 4-3. Physical and Chemical Properties of Technical PolybrominatedDiphenyl Ether (PBDE) MixturesProperty Pentabromodiphenyl ether Octabromodiphenyl ether Decabromodiphenyl etherMolecular weightColorPhysical stateMelting pointBoiling pointDensity (g/mL)OdorOdor threshold:WaterAirSolubility:WaterOrganic solvent(s)Partition coefficients:Log KowLog KocVapor pressureHenry’s Law constant(atm-m3/mole)AutoignitiontemperatureFlashpointFlammability limitsMixtureClear, amber to pale yellowa Highly viscous liquid-7 to -3°C (commercial)b >300°C (decomposition starts above 200°C)a,b 2.28 at 25°Ca; 2.25–2.28bNo dataNo dataNo data13.3 µg/L (commercial)b,d; 2.4 µg/L (pentabromodiphenyl ether component)b; 10.9 µg/L (tetrabromodiphenyl ether component)b10 g/kg methanol; miscible in toluened 6.64–6.97d; 6.57(commercial)b4.89–5.10e 2.2x10-7–5.5x10-7 mm Hg at 25°Cd; 3.5x10-7 mm Hg (commercial)b1.2x10-5g; 1.2x10-6e; 3.5x10-6f Decomposes above 200°Cb,d No dataNot applicable (flame retardant)b,d MixtureOff-whitea Powder85–89°C (commercial)c; 200°C (range, 167–257)a; 79–87°Ca; 170–220°Ca Decomposes at >330°C(commercial)c2.76a; 2.8 (commercial)cFainta No dataNo data<1 ppb at 25°C (commercial)c; 1.98 μg/L (heptabromodiphenylether component)c Acetone (20 g/L);benzene (200 g/L);methanol (2 g/L) all at25°Ca 6.29 (commercial)c5.92–6.22e 9.0x10-10–1.7x10-9 mmHg at 25°Cd; 4.9x10-8 mmHg at 21°C (commercial)c7.5x10-8e; 2.6x10-7e Decomposes above 330°C (commercial)cNo dataNot applicable (flame retardant)c959.22a Off-whitea Powdera 290–306°Ca Decomposes at >320,>400, and 425°Ca 3.0a; 3.25a Odorlessa Not applicableNot applicable<0.1 μg/LgNo data 6.265e6.80e 3.2x10-8 mm Hgf 1.62x10-6g; 1.93x10-8d; 1.2x10-8e; 4.4x10-8e Not applicablea NoneNon-flammablea PBDEs 3544. CHEMICAL AND PHYSICAL INFORMATIONTable 4-3. Physical and Chemical Properties of Technical PolybrominatedDiphenyl Ether (PBDE) MixturesProperty Pentabromodiphenyl ether Octabromodiphenyl ether Decabromodiphenyl etherConversion factors 1 ppm=23.48 mg/m3 at No data No data20°Cd Explosive limits Noneb,f Nonec No dataa WHO 1994a bENVIRON 2003a cENVIRON 2003bdEU 2001 eEstimated values were calculated using EPIWIN v4.10 (EPA 2014e).fHardy 2002agEstimated value was calculated using vapor pressure and water solubility values in table. PBDEs 3554. CHEMICAL AND PHYSICAL INFORMATIONTable 4-4. Physical and Chemical Properties of Some Polybrominated DiphenylEther (PBDE) CongenersCongenerVapor pressure (mm Hg)a,bWater solubility(μg/L)aHenry’s Law constant(atm m3/mol)a Log Kow c Log KOAdBDE 3 1.94x10-3 – – – –BDE 15 1.30x10-4 130 2.07254x10-4 – –BDE 17 – – – 5.74 9.30BDE 28 1.64x10-5 70 5.03331x10-5 5.94 9.50BDE 47 1.40x10-6 15 1.48038x10-5 6.81 10.53BDE 66 9.15x10-7 18 4.93461x10-6 – 10.82BDE 77 5.09x10-7 6 1.18431x10-5 – 10.87BDE 85 7.40x10-8 6 1.08562x10-6 – 11.66BDE 99 1.32x10-7 9 2.26992x10-6 7.32 11.31BDE 100 2.15x10-7 40 6.80977x10-7 7.24 11.13BDE 138 1.19x10-8 – – – –BDE 153 1.57x10-8 1 6.61238x10-7 7.90 11.82BDE 154 2.85x10-8 1 2.36862x10-6 7.82 11.92BDE 183 3.51x10-9 2 7.30323x10-8 8.27 11.96BDE 190 2.12x10-9 – – – –aTittlemier et al. 2002.bLiquid sub-cooled vapor pressures.cBraekevelt et al. 2003.dHarner 2001.– = no data reported; BDE = brominated diphenyl ether PBDEs 3564. CHEMICAL AND PHYSICAL INFORMATION“technical decaBDE,” or “technical DeBDE” which represents 97% BDE 209 congener with 3% nonaand octaBDE homolog impurities. The compositions of commercial product mixtures of PBDEs(e.g., technical penta-, octa-, and decaBDE) are given in Table 4-5. Trace analysis of these commercial mixtures for 15 different 2,3,7,8-substituted brominate dibenzop-dioxins and dibenzofurans revealed no detectable amounts of these substances (Hardy 2002a). Thecommercial decaBDE product has been analyzed for trace quantities of 15 2,3,7,8-substituted PBDDs and PBDFs. None of the analytes were present at or above the quantization limits established under an EPAtest rule (BFRIP 2002). While in today’s commercial PBDE samples, there are no measurable quantitiesof PBDDs/PBDFs, there are some materials that have reported quantifiable concentrations of these contaminants. For example, hexabromodibenzofurans have been detected in commercial decaBDEmixtures at concentrations as high as 200 μg/kg. In other PBDE mixtures (e.g., tetra- to hexaBDEs), thesum of tetra-, penta-, and hexabromodibenzofurans were reported at a concentrations of 8,000 μg/kg. Inaddition, tetra- and pentabromo-p-dibenzodioxins have been measured in commercial decaBDE atconcentrations of 0.05 and 0.35 μg/kg, respectively (WHO 1998). When pyrolyzed up to 900°C, PBDEs may produce PBDFs and PBDDs (Buser 1986; EU 2001). Theamount of PBDFs and PBDDs formed depends upon the conditions of pyrolysis. For example, 2,3,7,8-tetrabromodibenzofuran in ppm concentrations can be generated during pyrolysis of decaBDE in the temperature range of 400–700°C (Bieniek et al. 1989). PBDFs may also be produced during thepyrolysis of polymers containing PBDEs as flame retardants (Brenner and Knies 1993; Dumler et al. 1989, 1990; Lenoir et al. 1994). However, studies performed in the late 1980s may have suffered fromanalytical methods that could not differentiate between PBDDs/PBDFs formed (e.g., 2,3,7,8-substituted congeners) and decaBDE, which might have artificially elevated the concentrations of PBDEs detected(Hamm et al. 2001). PBDEs 3574. CHEMICAL AND PHYSICAL INFORMATIONTable 4-5. Typical Congener Composition of Penta-, Octa-, and Deca-Brominated Diphenyl Ether (BDE)Congener (weight percentage) PentaBDE OctaBDE DecaBDEBDE 47 25–37%BDE 99 35–50%BDE 100 6–10%BDE 153 3–5% 5–10%BDE 154 2–4% 1–5%BDE 183 40%BDE 196 8%BDE 197 21%BDE 203 5–35%BDE 206 2.2%BDE 207 7% 0.24%BDE 208 10% 0.06%BDE 209 97%Source: EPA 2010 PBDEs 3584. CHEMICAL AND PHYSICAL INFORMATIONThis page is intentionally blank. PBDEs 3595. PRODUCTION, IMPORT/EXPORT, USE, AND DISPOSAL5.1 PRODUCTIONThe commercial production of PBDEs generally involves bromination of diphenyl oxide to varyingdegrees. The degree of bromination is controlled either through stoichiometry or through control ofreaction kinetics (Pettigrew 1993). The commercial production of PBDEs began in late 1970s (WHO 1994a). There are no currentmanufacturers of technical PBDEs in the United States. About 98% of the global demand for thetechnical pentaBDE mixture resided in North America (Hale et al. 2003). PentaBDE and octaBDEmixtures were voluntarily withdrawn from the U.S. marketplace by their manufacturers at the end of2004, leaving only decaBDE being marketed for use in commercial products in the United States (EPA2010). In December of 2009, the two remaining U.S. producers of decaBDE, Albemarle Corporation and Chemtura Corporation (formerly known as the Great Lakes Chemical Corporation), and the largest U.S.importer, ICL Industrial Products, Inc., announced commitments to phase out manufacture and importation of decaBDE for most uses in the United States by December 31, 2012, and to end manufacture and import for all uses by the end of 2013 (EPA 2013j). In 2003, the EU passed a Directiveto ban the marketing and use of penta- and octaBDE that took effect in 2004. In 2008, the use ofdecaBDE was restricted by an EU Directive on the Restriction of the use of certain Hazardous Substances (RoHS) (EC 2014; EPA 2010).Table 5-1 lists the facilities in each state that manufactured or processed technical decaBDE in 2014, the intended use, and the range of maximum amounts of technical decaBDE that are stored on-site (TRI14 2016). The data from the Toxics Release Inventory (TRI) listed in Table 5-1 should be used with caution, however, since only certain types of facilities were required to report. The TRI is not an exhaustive list.Facilities are only required to report to the TRI if they manufacture or process more than 25,000 poundsof a TRI listed chemical during the year, or otherwise use more than 10,000 pounds, and have theequivalent of more than 10 full-time employees. According to the EPA, TRI data have certainlimitations. TRI data reflect releases and other waste management of chemicals, and not exposures of the public to those chemicals. TRI data alone are not sufficient to determine exposure or calculate potentialadverse effects on human health and the environment. PBDEs 3605. PRODUCTION, IMPORT/EXPORT, USE, AND DISPOSALTable 5-1. Facilities that Produce, Process, or Use Decabromodiphenyl EtherMinimum Maximum Number of amount on site amount on site Statea facilities in poundsb in poundsb Activities and usesc AR 2 1,000 99,999 7CA 1 10,000 99,999 2, 3, 7, 8, 10CT 1 10,000 99,999 8GA 3 1,000 99,999 7, 8IL 3 10,000 99,999 7IN 1 50,000,000 99,999,999 2, 3, 4, 8KS 1 1,000 9,999 8MA 2 10,000 999,999 7MS 1 10,000 99,999 7, 8NH 1 Not reported Not reported Not reportedNJ 1 10,000 99,999 7NV 1 10,000 99,999 7NY 1 1,000 9,999 7OH 4 1,000 99,999 2, 3, 7, 8PA 3 100,000 999,999 1, 3, 7, 8SC 5 10,000 99,999 7, 8TX 2 1,000 9,999 7VA 3 0 99,999 7aPost office state abbreviations used.bAmounts on site reported by facilities in each state.cActivities/Uses:1. Produce 6. Impurity 11. Chemical Processing Aid2. Import 7. Reactant 12. Manufacturing Aid 3. Onsite use/processing 8. Formulation Component 13. Ancillary/Other Uses4. Sale/Distribution 9. Article Component 14. Process Impurity5. Byproduct 10. RepackagingSource: TRI14 2016 (Data are from 2014) PBDEs 3615. PRODUCTION, IMPORT/EXPORT, USE, AND DISPOSAL5.2 IMPORT/EXPORTProduction and importation of the final commercial PBDE, decaBDE, ended on December 31, 2013. 5.3 USE PBDEs were used as additive flame retardants in thermoplastics. Additive flame retardants are physicallycombined with the polymer material being treated rather than chemically combined (as in reactive flame retardants). This means that there is a possibility that the flame retardant may diffuse out of the treatedmaterial to some extent. PBDEs were used in different resins, polymers, and substrates at levels rangingfrom 5 to 30% by weight (EU 2001). The commercial pentaBDE product was used predominantly (95–98%) for flame retardant purposes as anadditive in consumer products manufactured by the furniture industry (ENVIRON 2003a). It was used almost exclusively to flame retard flexible polyurethane foam (FPUF), which is used in bed mattresses and cushioning in upholstered products. The commercial pentaBDE was typically used in FPUF as an additive mixture with aromatic phosphate esters (e.g., mixture of 75% pentaBDE and 25% aromaticphosphate esters). Mattress FPUF contains approximately 2–3% flame retardant mixture and cushion FPUF contains 3–5% flame retardant mixture (ENVIRON 2003a). Scrap materials from both industrieshave been used as padding beneath carpets, and as a result, carpet padding likely contained 3–5% flame retardant mixture. However, not all of the FPUF found in cushion, mattress, and carpet padding productswere treated with commercial pentaBDE. Approximately 7.5% of the more than 2.1 billion pounds ofFPUF produced annually in the United States used the commercial pentaBDE product as a flame retardantadditive (ENVIRON 2003a). The majority of FPUF products treated with the commercial pentaBDEproduct were sold in California, the only state requiring by law that upholstered products achieve a prescribed level of ignition resistance (ENVIRON 2003a). A small percentage of pentaBDE was used incommercial adhesive products. Other former uses of commercial pentaBDE included coatings forspecialty textiles, printed circuit board components, hydraulic and oilfield completion fluids, and rubberproducts. In the past, automotive and airplane seating cushions contained FPUF with commercialpentaBDE. However, this use was discontinued in the early 1990s. Prior to approximately 1990, thecommercial pentaBDE product may have been used in small quantities as a flame retardant in specialtyfire-resistant clothing using polyurethane treatment and in polyurethane coatings in carpets (ENVIRON2003a). Commercial pentaBDE product was used in rigid polyurethane elastomers for instrumentcasings, and applied in printed circuit boards and microprocessor packaging previously (Betts 2006; PBDEs 3625. PRODUCTION, IMPORT/EXPORT, USE, AND DISPOSALHazrati and Harrad 2007). Electronic equipment containing pentaBDE produced in other countries(principally Asian) could also find its way into the United States (EU 2001). The commercial octaBDE was used by the plastics industry as an additive flame retardant formanufactured products. It was used almost exclusively to flame retard ABS terpolymers used in computer casings and monitors (ENVIRON 2003b). In the EU, approximately 95% of the totalcommercial octaBDE product sold to the electronics and plastics industries was used in ABS before it wasbanned (EU 2003a). Although data are not available in the United States, similar volumes were likely(ENVIRON 2003b). The commercial octaBDE product formerly used in ABS products was 12–18%weight loadings of flame retardant. OctaBDE was always used as a flame retardant in conjunction withantimony trioxide. Other minor uses for octaBDE, were high impact polystyrene (HIPS), polybutyleneterephthalate (PBT), and polyamide polymers (EU 2003a). Other former applications of octaBDEincluded use as additive flame retardant in polycarbonate, phenol-formaldehyde resins, and unsaturatedpolyesters (EU 2003a).The commercial decaBDE product was an additive flame retardant used in a variety of polymerapplications. Industry information indicates that decaBDE was used at loadings of 10–15% weight in polymers and always in conjunction with antimony trioxide (EU 2002). The major application fordecaBDE was in HIPS, which is used in the television industry for cabinet backs. It was also used for a large number of other polymers with end-uses in electrical and electronic equipment (e.g., computers, connectors, electrical boxes, wire, cable, etc.). Examples include polypropylene (for electronics), acetate copolymers (ethylene-vinyl acetate [EVA] and other copolymers for wire and cable), ethylene-propylenediene terpolymer (EPDM) and thermoplastic elastomers (for wire and cable), and polyester resins (forelectronics). Other minor uses included styrenic rubbers, polycarbonates, polyamides, and terephthalates,and small amounts are reported to be used in hot-melt adhesives (EU 2002). 5.4 DISPOSALPBDEs were used as flame retardants in a wide range of consumer products (see Section 5.3). In theUnited States, waste disposal of PBDE-containing consumer products is described as transfers to disposal(landfill), recycling, energy recovery (incineration), or publicly owned treatment works (POTWs)(Darnerud et al. 2001). No other information was located on the past or present volumes of PBDE-containing consumer products disposed of by each method of waste transfer. PBDEs 3635. PRODUCTION, IMPORT/EXPORT, USE, AND DISPOSALLandfill disposal of plastic consumables containing pentaBDE (e.g., polyurethane foams), octaBDE (e.g., computer monitors), and decaBDE (e.g., televisions) to landfills is likely to increase in the United States due to their limited useful lifespan. Given that all PBDEs have low water solubility (see Table 4-4), thepotential for leaching of PBDE from landfills appears to be small (EU 2002). Well-designed landfills willinclude measures to minimize leaching and those measures would also be effective in minimizing the leaching of any PBDE particulates present (EU 2002). PBDEs have been detected in landfill leachate andlandfill related aqueous samples worldwide (Daso et al. 2013; Kwan et al. 2013; Odusanya et al. 2009;Oliaei et al. 2010; Stubbings and Harrad 2014). The presence of hydrophobic compounds like PBDEs in leachate is expected to be a result of enhanced leachability due to the presence of other constituents present in the leachate (Stubbings and Harrad 2014). Mass transfer evaluation of PBDEs from e-waste solids found that lower pH conditions resulted in higher transfer of PBDEs to the aqueous phase, with thehighest concentration of PBDEs detected at pH 5 and 25°C (Danon-Schaffer et al. 2013). The levels ofPBDEs in the aqueous phase did show a trend with temperature at the temperature range evaluated, 10– 25°C.Incineration of waste materials containing PBDEs is thought to be a potential source of PBDFs and/orPBDDs. The formation of PBDFs/PBDDs as a result of uncontrolled landfill fires is also a possibility,although no data are available on the scale of this source. The results of pyrolysis experiments showed that PBDEs can form PBDFs and PBDDs (in much smaller quantities) under a wide range of heatingconditions. If chlorine is present, mixed halogenated furans/dioxins can be formed (Oberg et al. 1987;Zier et al. 1991). Unless sufficiently high temperatures and long residence times are maintained, PBDFs/PBDDs can be generated during the incineration of products containing PBDEs. When heavymetals are present, the concentration of PBDDs and PBDFs are higher than when no metals are present.Sakai et al. (2001) measured residues of PBDFs/PBDDs in effluents from a municipal incineration plantburning domestic waste materials. Flue gases, fly ash, and bottom ash reportedly contained PBDFs/PBDDs at concentration ranges of 0.28–3.3 ng/N m3, 0.082–13 ng/g, and 0.0058–27 ng/g, respectively. However, modern, properly operated municipal waste incineration should not emitsignificant quantities of PBDFs/PBDDs, regardless of the composition of municipal waste (WHO 1994a). In the United States, waste disposal of industrial by-products containing PBDEs may also be described as transfers to disposal (landfill), recycling, energy recovery (incineration), industrial treatment works, orPOTWs. The types of waste transfer may be different for manufacturing versus processing sectors, andalso from within different types of processing. Waste disposal from manufacturing processes is predominantly to secure chemical landfills (e.g., those built with liners and leachate collection). Plastic PBDEs 3645. PRODUCTION, IMPORT/EXPORT, USE, AND DISPOSALprocessors typically transfer most waste to disposal (landfill), recycling, energy recovery (incineration), and industrial treatment works, while minimal releases are to POTWs. In contrast, textile processors typically transfer most waste to POTWs. This difference in waste transfers between the plastic and textile sectors is because textile processors use water in their processing operation and other processors (e.g.,processors of plastic) do not. Recycling of plastic materials containing PBDEs is a common practice in industry. It has beendemonstrated that decaBDE-containing resins can be successively recycled without generation ofPBDDs/PBDFs (Brenner and Knies 1990; Donnelly et al. 1987; McAllister et al. 1990). For example, virgin HIPS resins (containing antimony trioxide [Sb2O3] and decaBDE) and repeatedly ground and injected molded (e.g., “recycled”) HIPS/decaBDE/Sb2O3 resins both met the requirements of the GermanChemicals Banning Ordinance with respect to 2,3,7,8-substituted PBDD/PCDF congeners. These resins were at least 1 order of magnitude below the regulated limit values for PBDDs/PCDFs (1 ppb for the sumof four congeners, 5 ppb for the sum of all eight regulated congeners). PBDEs 3656. POTENTIAL FOR HUMAN EXPOSURE6.1 OVERVIEWPBDEs have not been identified in any of the 1,832 hazardous waste sites that have been proposed forinclusion on the EPA National Priorities List (NPL) (ATSDR 2015). However, the number of sitesevaluated for PBDEs is not known. The widespread use of PBDEs since the late 1970s has resulted in their presence in the environment. PBDEs were released into the environment from their manufacture and use as additive flame retardants inthermoplastics in a wide range of products (WHO 1994a). Waste containing PBDEs may be incinerated as municipal waste, deposited in landfills, discharged to municipal sewage-treatment plants, or emitteddirectly to the atmosphere as particulates (Darnerud et al. 2001). Adsorption of PBDEs generally increases as bromination of PBDEs and organic carbon content of soiland sediment increase. As a result, most PBDEs have little or no mobility in soil and are not expected to leach (e.g., into groundwater). PBDEs, particularly lower BDE homologs (e.g., tri- and tetraBDE), have the potential for long-range transport in the atmosphere (Dodder et al. 2000). The detection of PBDEs in remote regions of the world suggests that long-range transport of these congeners is occurring (Dickhut etal. 2012; Hung et al. 2010). Biodegradation is a slow environmental fate process for PBDEs, but undercertain conditions, some PBDEs compounds (e.g., decaBDE) may degrade by direct photolysis to formlower-brominated congeners. However, determining the rate and extent of degradation processes (e.g., biodegradation and photolysis) for PBDEs, such as decaBDE and pentaBDE commercial mixtures, is stillan active area of research.Studies of the biota indicate that lower-brominated congeners (e.g., BDE 47) are being preferentiallybioconcentrated. Lower-brominated diphenyl ether (e.g., tetra- and pentaBDE) concentrations increase with respect to trophic level; thus, organisms that reside higher on food chains tend to have higherconcentrations of these brominated diphenyl ethers (Shaw et al. 2009). Body-burden data indicate thatthe general population is exposed to PBDEs through a variety of pathways (CDC 2015; Lorber 2008;Trudel et al. 2011). The primary exposure pathway to PBDEs for residents of North America is through indoor dust contact (ingestion and dermal exposure) (EPA 2010; Lorber 2008. Dust contact is also the primary exposure pathway for BDE 209 in the United Kingdom. For Europeans, food consumption appears to be the primary exposure pathway for most congeners (Abdallah and Harrad 2014; Law et al.2008; Trudel et al. 2011). Body burden data, as well as intake modeling, suggest that infants and toddlers PBDEs 3666. POTENTIAL FOR HUMAN EXPOSUREhave higher exposures to PBDEs as compared to older children or adults. PBDE levels increase frominfant to toddler and then PBDE concentrations gradually decrease at older ages. Most studies indicate that concentrations of PBDEs in body fluids and tissues are a factor of 10–100-fold higher for individualsliving in the United States compared to individuals living in other regions of the world (e.g., Europe). 6.2 RELEASES TO THE ENVIRONMENTThe Toxics Release Inventory (TRI) data should be used with caution because only certain types offacilities are required to report (EPA 2005). This is not an exhaustive list. Manufacturing and processing facilities are required to report information to the TRI only if they employ 10 or more full-timeemployees; if their facility is included in Standard Industrial Classification (SIC) Codes 10 (except 1011, 1081, and 1094), 12 (except 1241), 20–39, 4911 (limited to facilities that combust coal and/or oil for thepurpose of generating electricity for distribution in commerce), 4931 (limited to facilities that combustcoal and/or oil for the purpose of generating electricity for distribution in commerce), 4939 (limited to facilities that combust coal and/or oil for the purpose of generating electricity for distribution incommerce), 4953 (limited to facilities regulated under RCRA Subtitle C, 42 U.S.C. section 6921 et seq.), 5169, 5171, and 7389 (limited S.C. section 6921 et seq.), 5169, 5171, and 7389 (limited to facilitiesprimarily engaged in solvents recovery services on a contract or fee basis); and if their facility produces,imports, or processes ≥25,000 pounds of any TRI chemical or otherwise uses >10,000 pounds of a TRIchemical in a calendar year (EPA 2005).The widespread use of PBDEs from the 1970s until 2013 resulted in their presence in the environment. PBDEs were released into the environment from their manufacture and use in a wide range of consumerproducts (WHO 1994a). PBDEs were used as additive flame retardants in thermoplastics. Additive flame retardants are physically, rather than chemically, combined with polymers. Thus, there is a possibility that some PBDEs congeners may diffuse out of the treated materials to some extent (EU2001). Although these substances are no longer manufactured in the United States and Europe, thedisposal of consumer products that contain penta-, octa-, and decaBDE will result in their continued release to the environment. Waste from products containing PBDEs may be incinerated as municipalwaste, deposited in landfills, or discharged to municipal sewage-treatment plants (Darnerud et al. 2001).6.2.1 AirEstimated releases of 2,180 pounds (~0.9 metric tons) of decaBDE to the atmosphere from 34 domesticmanufacturing and processing facilities in 2014, accounted for about 1.5% of the estimated total PBDEs 3676. POTENTIAL FOR HUMAN EXPOSUREenvironmental releases from facilities required to report to the TRI (TRI14 2016). These releases are summarized in Table 6-1. There are no TRI data for penta- or octaBDE.The estimated release of decaBDE from the 2014 TRI continues to reflect a decreasing trend as production slowed and was eventually discontinued in 2013. The total on-site and off-site releases ofdecaBDE since 1998 are illustrated in Figure 6-1 (TRI14 2016). No quantitative information was located on the releases of the pentaBDE technical mixtures to the atmosphere from its former production and use. However, the release of pentaBDE technical mixtures toair had the potential to occur during the curing phase, since the polyurethane foam was at elevatedtemperatures (e.g., up to 160°C) for several hours during this phase. Since pentaBDE technical mixtureswere additive flame retardants, they are subject to volatilization or leaching from the polymer matrix during the lifetime of the use of the foam article. Losses of foam particles containing the substance (e.g.,due to abrasion) may also occur. However, most congeners in pentaBDE technical mixtures have verylow vapor pressures (see Table 4-3) and therefore, losses from polyurethane foam due to volatilization would be expected to be low. Migration of pentaBDE technical mixtures from consumer products may bea significant diffuse source of lower-brominated congeners of pentaBDE technical mixtures to theatmosphere. Although no studies were found that determined the migration rate of pentaBDE technicalmixtures from polymers into the air, estimates have been made. The estimated migration rate forpentaBDE technical mixtures is 0.39% per year (Danish EPA 1999).Similarly, no quantitative information is available on emissions of octaBDE technical mixtures to theatmosphere from production operations. The major sources of air emissions of octaBDE technicalmixtures were thought to be a result of grinding and bagging operations. The EPA National Center for Environmental Assessment (NCEA), Office of Research and Developmentcompleted a comprehensive exposure assessment of PBDEs (EPA 2010). A series of studies weresummarized that estimated the release of PBDEs from various products under laboratory conditions. Twocomputer workstations manufactured after 2000 consisting of a monitor, computer, keyboard, mouse, and printer were used for 93 and 150 days and PBDE concentrations were monitored during their operation. BDE 47, BDE 100, BDE 99, and BDE 85 concentrations in surrounding air were <0.3 ng/m3 for one ofthe workstations; however, concentrations of BDE 47, BDE 100, and BDE 99 were 150, 28, and 61 ng/m3, respectively, in air monitored for the second workstation. An emission test was summarizedthat used the back panel of a television set treated with octaBDE manufactured before 1979. Maximum PBDEs 3686. POTENTIAL FOR HUMAN EXPOSURETable 6-1. Releases to the Environment from Facilities that Produce, Process, orUse Decabromodiphenyl EtheraReported amounts released in pounds per yearb Total releaseStatec RFd Aire Waterf UIg Landh Otheri On-sitej Off-sitek On- and off-siteAR 2 120 0 0 5,374 0 120 5,374 5,494CA 1 1 0 0 139 0 1 139 140CT 1 0 0 0 0 1,747 0 1,747 1,747GA 3 347 0 0 2,917 0 347 2,917 3,264IL 3 0 0 0 5 0 0 5 5IN 1 5 0 0 0 0 5 0 5KS 1 NR NR NR NR NR NR NR NRMA 2 18 5 0 84 1 23 85 108MS 1 28 0 0 0 0 28 0 28NH 1 NR NR NR NR NR NR NR NRNJ 1 0 0 0 0 150 0 150 150NV 1 NR NR NR NR NR NR NR NRNY 1 0 0 0 137 0 0 137 137OH 4 505 0 0 75 112,501 505 112,576 113,081PA 3 1,152 0 0 6,919 0 1,152 6,919 8,070SC 5 0 0 0 67 0 0 67 67TX 2 5 0 0 6,182 0 5 6,182 6,187VA 1 0 0 0 2,504 0 0 2,504 2,504Total 34 2,180 5 0 24,403 114,399 2,185 138,802 140,987aThe TRI data should be used with caution since only certain types of facilities are required to report. This is not an exhaustive list. Data are rounded to nearest whole number.bData in TRI are maximum amounts released by each facility.cPost office state abbreviations are used.dNumber of reporting facilities.eThe sum of fugitive and point source releases are included in releases to air by a given facility.fSurface water discharges, waste water treatment-(metals only), and publicly owned treatment works (POTWs) (metaland metal compounds).gClass I wells, Class II-V wells, and underground injection.hResource Conservation and Recovery Act (RCRA) subtitle C landfills; other onsite landfills, land treatment, surface impoundments, other land disposal, other landfills.iStorage only, solidification/stabilization (metals only), other off-site management, transfers to waste broker fordisposal, unknownjThe sum of all releases of the chemical to air, land, water, and underground injection wells.kTotal amount of chemical transferred off-site, including to POTWs.NR = not reported; RF = reporting facilities; UI = underground injectionSource: TRI14 2016 (Data are from 2014) PBDEs 3696. POTENTIAL FOR HUMAN EXPOSUREFigure 6-1. Total On- and Off-Site Releases of Decabromodiphenyl Ether, 1998–2014Source: TRI14 (2016) PBDEs 3706. POTENTIAL FOR HUMAN EXPOSUREconcentrations of BDE 28, BDE 47, BDE 66, BDE 100, and BDE 99 were reported as 0.5, 8, 0.24, 0.27, and 0.84 ng/m3, respectively (EPA 2010). These data suggest that older treated consumer products maycontinue to release PBDEs long after they were originally treated.More generalized approaches were summarized that could be used to estimate the possible totalvolatilization of PBDEs from treated plastic products (EPA 2010). The EU utilized a regression derived equation to estimate the percentage loss of PBDE that volatilizes from plastic components treated with PBDEs:PercentVolatilized =1.1×106 ×VP × SL where VP is the vapor pressure in units of mm Hg and SL is theservice life of the product, assumed to be 10 yearsFor instance, using a vapor pressure of 3.47x10-8 mm Hg for decaBDE, the volatilization loss after10 years would be approximately 0.38% (EPA 2010). Since approximately 6,710 metric tons ofdecaBDE were used in plastics the EU before it was banned, the total loss to air over the assumed 10-yearlifetime would be approximately 25.5 metric tons (EPA 2010). Breivik et al. (2002) developed a regression-derived equation using the octanol-air partition coefficient (KOA) to estimate emission factors of PCBs from commercial sealants, which was also applied to estimate the emission factors of PBDEslog EF = −0.839 × log KOA + 4.83 The emission factor (EF) is the ratio of the mass of PBDE emitted divided by the mass PBDE used peryear. Both equations above were used to estimate total emissions of penta-, octa-, and decaBDE fromproducts used in the United States (EPA 2010). 6.2.2 WaterEstimated releases of 5 pounds (~0.002 metric tons) of decaBDE to surface water from 34 domestic manufacturing and processing facilities in 2014, accounted for <1% of the estimated total environmentalreleases from facilities required to report to the TRI (TRI14 2016). These releases are summarized inTable 6-1. PBDEs 3716. POTENTIAL FOR HUMAN EXPOSUREIndustrial and urban effluents are sources of PBDEs to surface waters and sediments. Limited data onindustrial and urban effluents were located for the United States. Hale et al. (2002) measured the concentration of PBDEs in soil and stream sediments collected near a polyurethane manufacturing plant(near the Dan River, Virginia). Summed concentrations of BDE 47, BDE 99, and BDE 100, the dominantcongeners in these samples, ranged from <1 to 132 µg/kg (ng/g) dry weight. In 1995, sediment sampleswere collected up- and downstream near an area where the Swedish plastics industry used brominatedflame retardants (Sellström and Jansson 1995; Sellström et al. 1998a). Samples were analyzed for tetraBDEs (50 ng/g dry weight) and pentaBDEs (sum of three congeners, 2,300 ng/g dry weight). ThesePBDEs were found in higher concentrations downstream of the plant than upstream, which indicates thatthe plastics industry was the most likely source of these compounds. Surficial sediment samples were collected at eight locations along River Viskan near several textile manufacturing facilities that usedvarious brominated flame retardants in the production of textiles. The concentrations of BDE 47,BDE 99, BDE 100, and BDE 209 in sediments increased as samples were collected further downstreamwhere additional industries were located (Sellström et al. 1998a). The lowest concentrations of PBDEswere found upstream of the textile industries. The combined concentration of BDE 47, BDE 99, andBDE 100 ranged from not detected to 120 ng/g (µg/kg) dry weight; the concentration of BDE 209 ranged from not detected to 16,000 ng/g (µg/kg) dry weight. Allchin et al. (1999) surveyed the concentrations ofPBDEs in sediments from several rivers and estuaries in Great Britain. Sediments were collectedupstream and downstream of suspected sources of pentaBDE and octaBDE, including a manufacturer, several industries, landfills, and a reference site. The highest concentrations of BDE 47, BDE 99, pentaBDE (as BDE 71), and octaBDE (as BDE 79) were in sediments near or downstream from a manufacturing site at Newton Aycliffe in River Skerne. The highest concentrations of decaBDE (as BDE 83) were found downstream of a sewage-treatment plant on River Calder. High concentrations were also detected on River Skerne downstream of a manufacturing site. BDE 99 concentrations were identicalor slightly higher than BDE 47 in most sediments (Allchin et al. 1999). The sum of five pentaBDEcongeners (BDE 47, BDE 99, BDE 100, BDE 153, and BDE 209) ranged from 0.07 to 10.6 ng/g (μg/kg)dry weight in freshwater sediments from Denmark (Christensen and Platz 2001). The highestconcentrations were found in sediment close to populated areas.A study conducted by the U.S. Geological Survey (USGS) analyzed waste water treatment plant (WWTP)effluent from nine cities located in Oregon and Washington for anthropogenic compounds, includingPBDEs (USGS 2012). Detectable levels in the low ng/L range were observed at every WWTP, and the highest concentrations measured were for congeners BDE 47, BDE 99, and BDE 100. The greatest PBDEs 3726. POTENTIAL FOR HUMAN EXPOSUREPBDE concentrations were observed in Richland and Portland, Oregon. The Portland PBDE valuesshowed varying concentrations as a function of the time of the day that samples were obtained. The lowest PBDE levels were observed in the morning hours and afternoon and then increased 2–4 times byevening hours. Shreder and LaGuardia (2014) measured PBDEs in the effluent of laundry waste waterfrom 20 residences located near the Columbia River in Washington state. BDE 47 and BDE 209 weredetected in the laundry waste water effluent of all 20 homes at median levels of 1,230 and 140 ng/L, respectively. The median concentration of total PBDE (sum of BDE 28, 47, 66, 85, 99, 100, 153, 154, 183, 206, and 209) in the laundry waste water was reported as 2,550 ng/L. BDE 47, BDE 49, and BDE 209 were also detected in the influents of two WWTPs near the Columbia River, Washington thatprimarily serve residential households. The sum total levels of BDE 47, BDE 49, and BDE 209 in theinfluents of the WWTPs ranged from 35 to 206 ng/L. Effluent PBDE levels were below the detectionlimit at one WWTP and 28.2 ng/L at the other facility (Shreder and LaGuardia 2014). Although the available information indicates that leaching of PBDEs from landfills is minimal, movementof polymer particles containing pentaBDE, octaBDE, and decaBDE commercial mixtures within thelandfill could lead to entry into leachate water of groundwater. PBDEs have been detected in landfillleachate and landfill related aqueous samples (Daso et al. 2013; Kwan et al. 2013; Odusanya et al. 2009;Oliaei et al. 2010; Stubbings and Harrad 2014). The presence of hydrophobic compounds like PBDEs in leachate is expected to be a result of enhanced leachability due to the presence of other constituents present in the leachate (Stubbings and Harrad 2014). Mass transfer evaluation of PBDEs from e-waste found that lower pH conditions resulted in higher transfer of PBDEs to the aqueous phase, with thehighest concentration of PBDEs detected at pH 5 (Danon-Schaffer et al. 2013). It is not currentlypossible to assess the significance of this type of process. Well-designed landfills already include measures to minimize leaching in general, and these measures would also be effective in minimizingleaching of any PBDEs present (EU 2002, 2003a).6.2.3 Soil Estimated releases of 24,403 pounds (~11.1 metric tons) of decaBDE to soils from 34 domesticmanufacturing and processing facilities in 2014, accounted for about 17% of the estimated totalenvironmental releases from facilities required to report to the TRI (TRI14 2016). These releases aresummarized in Table 6-1. PBDEs 3736. POTENTIAL FOR HUMAN EXPOSUREPBDEs are released to land (i.e., landfills) as waste from their manufacture (both raw material and polymer) and as municipal wastes with the disposal of consumer products. The disposal of consumerproducts containing PBDEs is likely to increase worldwide due to rapid obsolescence of plastic products.PBDEs may be present in biosolids and may therefore be inadvertently released to soils from the use ofbiosolids as a nutrient amendment to agricultural soils. Biosolids are sewage sludge that has been treatedto meet regulatory requirements for land application and must adhere to concentration limits and loadingrates for chemical pollutants, treatment and use requirements for controlling and reducing pathogens and the attraction of vectors, and management practices (NRC 2002). PBDEs were detected in biosolidsdestined for land applications in four different regions of the United States (Hale et al. 2001c). The totalconcentrations of pentaBDE in biosolids ranged from 1,100 to 2,290 μg/kg dry weight. The concentration of decaBDE (BDE 209) varied widely among biosolids from different regions; the concentration ofBDE 209 ranged from 84.8 to 4,890 μg/kg dry weight in the biosolid samples. Kim et al. (2013b)analyzed 288 samples of sludge and biosolids from 15 WWTPs in Canada. Total PBDE levels were 230– 82,000, 530–8,800, and 420–6,000 μg/kg, in primary sludge, waste biological sludge and treated biosolids respectively. BDE 209, BDE 99, and BDE 47 were reported as the predominant congeners. In thebiosolids, these three congeners accounted for approximately 80% of the total. The median percentagesof BDE 209, BDE 99, and BDE 47 were 47, 16, and 16%, respectively, of the total amount of all PBDEcongeners in the biosolids.6.3 ENVIRONMENTAL FATE6.3.1 Transport and Partitioning PBDEs exist in both the vapor phase and the particulate phase in the atmosphere. Particulate-phase PBDEs will be removed from the atmosphere by wet and dry deposition. A vapor phase–particulate phase analysis of indoor air samples obtained from Birmingham, United Kingdom found that 66–86% ofBDE 47, 54–65% of BDE 99, 63–74% of BDE 100, <20–48% of BDE 153, and 37–48% of BDE 154 existed in the vapor phase (Harrad et al. 2004). Strandberg et al. (2001) performed a vapor phase– particulate phase analysis of outdoor air samples obtained from the Great Lakes region and found thatabout 80% of BDE 47, 55–65% of BDE 100 and BDE 99, and 30% of BDE 154 and BDE 153 existed in the gas phase. Several PBDE congeners have been detected in Arctic regions, suggesting that thesesubstances undergo aerosol-mediated, long-range atmospheric transport. BDE 47, BDE 99, BDE 100, and BDE 209 were measured in air, snow, and sea ice samples throughout western Antarctica between 2001 and 2007 (Dickhut et al. 2012). Fourteen PBDE congeners are monitored for, and have been PBDEs 3746. POTENTIAL FOR HUMAN EXPOSUREdetected at, Alert and Nuuk monitoring stations as part of the Arctic Monitoring and AssessmentProgramme (Hung et al. 2010), indicating the importance of long-range transport as an environmental fateprocess for these substances. In water, PBDEs are expected to adsorb strongly to suspended solids and sediment, and bioconcentrate in aquatic organisms. The volatilization of PBDEs from water to air is expected to be attenuated byadsorption in the water column. In soil, PBDEs are adsorbed strongly and will be immobile. They arenot likely to leach into groundwater. Volatilization of PBDEs from soil to air is limited by the lowvolatility of PBDEs and strong adsorption of PBDEs to soil. There is potential for PBDEs to volatilizefrom soil to air, particularly if the organic carbon content of the soil is low, as demonstrated by PBDEsbeing monitored in air as described in Section 6.4.1.PBDEs adsorb strongly onto suspended solids and sediments in the water column. Volatilization ofPBDEs from water surfaces will be attenuated by adsorption, and is thus not an important fate process. Sediment-water partition coefficients (Kp) have been measured for several components of commercialpentaBDEs (Watanabe 1988). Kp values for tetra-, penta-, and hexaBDEs are 28,300, 49,200, and 62,700 L/kg, respectively, which suggest strong partitioning to sediment. Log organic carbon-waterpartition coefficients (Koc) were estimated for PBDEs: di- (4.11); tri- (4.35–4.41); tetra- (4.57–4.73);penta- (4.89–5.17); hexa- (5.11–5.69); octa- (5.92–6.22); and deca- (6.80) (Lyman et al. 1990).DecaBDE and octaBDE commercial products do not bioconcentrate in fish to the same extent ascongeners from the penta mixture. Monitoring data show that higher-brominated congeners such as BDE209 are taken up in marine organisms. The reported bioconcentration factors (BCFs) for commercialdecaBDE mixtures are typically <50 (Hardy 2002b). A single study on a mixed range of PBDEs, between hexaBDE and decaBDE, indicated little bioconcentration in carp (e.g., Cyprinus carpio) with aBCF of <4 after 8 weeks of exposure (WHO 1994a). A bioconcentration study was carried out with rainbow trout under static conditions. The concentration of 14C-decaBDE/L in water was 20 μg. Fish were exposed to decaBDE for 0, 0.5, 1, 2, 4, 6, 12, 24, or 48 hours. For each of the exposure periods, there was no measurable accumulation of decaBDE in flesh, skin, or viscera (WHO 1994a). Thebioconcentration of BDE 209 was studied by exposing zebrafish embryos to BDE 209 at concentrationsof 0, 0.08, 0.38, and 1.92 mg/L until 14 days post-fertilization (Chen et al. 2012). BCFs of 29, 9, and 20 were calculated for exposure of 0.08, 0.38, and 1.92 mg/L, respectively. Several lower-brominated congeners were also detected in the larvae, with the main metabolite being nonaBDE. These results are PBDEs 3756. POTENTIAL FOR HUMAN EXPOSUREconsistent with other finding that indicate that BDE 209 was bioavailable and taken up by zebrafish larvaefrom spiked sediments (Garcia-Reyero et al. 2014). An abundance of monitoring data illustrates the uptake of lower-brominated diphenyl ethers by aquaticorganisms, which results in bioconcentration (see Section 6.4.4). The commercial pentaBDE productundergoes bioconcentration with a BCF of approximately 14,000 (Hardy 2002b). Congener componentsof pentaBDE commercial product bioconcentrate to different extents. For example, approximately 50– 70% of PBDEs detected in fish is a single isomer (BDE 47). The next most prominent isomer is typicallyBDE 99 followed by BDE 100. In a laboratory study of Baltic blue mussels (Mytilus edulis L), BCFsfrom water absorption were found to be 1,300,000 for BDE 47, 1,400,000 for BDE 99, and 1,300,000 for BDE 153 (Gustafsson et al. 1999). At several sites along the coast and in the Schelde Estuary (the Netherlands), BCFs for blue mussels were determined (Booij et al. 2000). The maximum BCFs were1x109 for BDE 99 and BDE 100, ≈2.5x107 for BDE 28, ≈2.5x108 for BDE 47 and ≈1.6x108 for BDE 153. Biomagnification of PBDE congeners BDE 47, BDE 99, BDE 100, BDE 153, BDE 154, and BDE 155 inthe marine food chain was demonstrated by comparing concentrations in the blubber of harbor seals with their prey fish (Shaw et al. 2009). Biomagnification factors (BMFs) from fish to seals were 21.4–109, 17.9–213, 6.9–29.8, 148–700, 11.3–447, and 12.4–236 for BDE 47, BDE 99, BDE 100, BDE 153, BDE154, and BDE 155, respectively (Shaw et al. 2009). BDE 209 was detected at measurable concentrationsin fish and seal tissue, although it did not appear to biomagnify like the other congeners. A laboratorystudy was conducted using juvenile carp fed BDE 209 amended food over a 60-day period (Stapleton etal. 2004). BDE 209 was not highly accumulated in the carp; however, seven debrominated lowercongeners not initially present (penta- to octaBDEs) were detected in whole fish samples and liver tissue,suggesting that while BDE 209 did not accumulate in the fish, it may be a source of lower-brominated metabolites in aquatic organisms. Other studies have demonstrated the biomagnification of lower-brominated PBDE congeners. Haglund etal. (1997) examined the concentrations of tetra- to hexaBDEs in herring, salmon muscle, and gray and ringed seals collected along the Swedish coast of the Baltic Sea between 1981 and 1988. Theconcentrations of tetraBDEs (e.g., BDE 47) were found to increase with trophic level. Concentrations ofPBDEs in herring and their predators, grey seal and guillemot, all collected at the same location of the Baltic Sea, have been compared to estimate potential biomagnification (de Wit 2002). The herring werecaught in the autumn of the same year as guillemot egg collection (1987). BMFs for guillemot egg versus herring were 19, 17, and 7.1 for BDE 47, BDE 99, and BDE 100, respectively. Burreau et al. (2000)analyzed small herring and salmon from the Atlantic Ocean (near Iceland) for several PBDEs. The PBDEs 3766. POTENTIAL FOR HUMAN EXPOSUREcalculated biomagnification factors for Atlantic salmon versus small herring were 3.5, 3.8, and 6.0 for BDE 47, BDE 99, and BDE 100, respectively. These authors concluded that biomagnification wasoccurring for the lower-brominated congeners.Biosolids from the Metropolitan Water Reclamation District of Greater Chicago, Stickney WWTP,collected between 2004 and 2007, were applied at two sites at a depth of 15−20 cm (Hale et al. 2012). Maximum total soil PBDE concentrations were 565 and 1,810 μg/kg for high clay soil and sandy soil,respectively. Corn grown at the two sites after the third year of annual biosolid application was evaluated for PBDEs using GC/MS with ENCI. PBDEs were not detected in the 46 grain, stover, or root samplesexamined, suggesting little uptake by corn from soils amended using biosolids. However, earthwormsexposed to PBDE containing biosolids were shown to accumulate these substances (Gaylor et al. 2013). Earthworms were exposed to a Class B anaerobically digested biosolid (ADB) containing5,560±440 μg/kg dry weight total penta congeners (BDE 47, 99, 100, 153, 154, and 183) and a composted biosolid (CB) containing 1,130±79 μg/kg dry weight total penta congeners over the course of a 28-day incubation period. Total penta PBDE body burdens in worms exposed to ADB amended soils were about5 times greater than those in the substrate, and worms exposed to CB amended soils had body burdensabout 4 times greater than in the substrate.6.3.2 Transformation and DegradationPhotolysis appears to be the dominant transformation process for PBDEs. However, the importance ofphotochemical transformation reactions in the environment cannot be determined due to lack ofinformation. Based on a very limited number of studies, biodegradation does not appear to be significantfor PBDEs.6.3.2.1 AirIn air, PBDEs may undergo indirect photolysis with hydroxyl radicals or direct photolysis with sunlight. Vapor-phase PBDEs may be degraded in the atmosphere by reaction with photochemically producedhydroxyl radicals. The half-lives for this reaction in air are estimated to be 29, 140, and 476 days, respectively, for penta-, octa-, and decaBDE homologs, calculated using a structure estimation method (Meylan and Howard 1993). This estimation is calculated using an atmospheric concentration of5x105 hydroxyl radicals per cm3 and is based on a 24-hour day of sunlight. The half-lives of PBDEs that are expected to be present in the particulate phase in the air will be longer than the estimated half-lives PBDEs 3776. POTENTIAL FOR HUMAN EXPOSUREcalculated for the gas-phase reaction. Thus, for the higher-brominated PBDEs (e.g., octa- and decaBDEs), indirect photolysis with hydroxyl radicals will be less important. In water, some PBDEs have been reported to undergo direct photolysis (Hua et al. 2003). Likewise,PBDEs present in the vapor phase (e.g., tetraBDE) or as particulates (e.g., decaBDE) may also undergophotolysis in the atmosphere. However, the rate and extent of the photolysis of PBDEs in air cannot beevaluated due the lack of information.6.3.2.2 WaterPBDEs absorb light in the environmental spectrum. Hua et al. (2003) found that decaBDE and thecommercial octaBDE absorbed light up to 325 nm, which indicates that these compounds may besusceptible to photodegradation at environmental wavelengths (Hua et al. 2003). Di- and tetraBDEs were reported to absorb minimal light at wavelengths >300 nm. This trend suggests that the lower-brominated diphenyl ethers (e.g., pentaBDE commercial mixtures) will be less susceptible to photolysis compared tooctaBDE and decaBDE commercial mixtures.PBDEs undergo debromination by direct photolysis in organic solvents and organic solvent:watermixtures. Laboratory studies of the photolytic breakdown of decaBDE in toluene have shown that it isdebrominated by ultraviolet (UV) light to hexaBDE and that photolysis occurs very rapidly (Sellström etal. 1998b). The photolysis half-life in toluene was <15 minutes. However, the amounts of lowerbrominated congeners appear to be small (EU 2002). The photolysis of decaBDE (and tetra-, penta-, hexa-, hepta-, and octaBDEs) was reported in an 80:20 mixture of methanol:water at wavelengths>290 nm (EU 2002). The rate of photodegradation was found to increase with increasing degree ofbromination. DecaBDE was found to degrade with a half-life of around 30 minutes, while half-lives fortetra-, penta-, hexa-, hepta-, and octaBDEs were 12–16 days, 2.4 days, 1.2 days, 1.2 days, and 5 hours, respectively. The decomposition products of decaBDE were identified to be PBDEs (with >6 bromineatoms per molecule) and polybrominated furans (with <6 bromine atoms per molecule). Results of thisstudy indicate that the photochemical stability of PBDEs increases with decreasing bromination (EU2002). Rayne et al. (2003b) reported that BDE 15 photodegraded in organic (acetonitrile-methanol) andaqueous (H2O:acetonitrile; 1:1 v/v) solvent systems at a wavelength of 300 nm. Reductive brominationwas reported to be much slower in the aqueous system (e.g., 73% remained after 300 minutes) compared to the organic system (where 51 and 41% remained after 30 minutes). However, these studies were conducted in the presence of organic solvents, which are not representative of conditions found in the PBDEs 3786. POTENTIAL FOR HUMAN EXPOSUREenvironment. Organic solvents can act as hydrogen donors in photolysis reactions, which will potentiallyaffect the distribution of products formed.The photolysis of PBDEs was examined under environmentally relevant conditions. Hua et al. (2003)studied the degradation of decaBDE in several different experiments: (1) on humic acid-coated silica particles; (2) on glass surfaces in contact with aqueous humic acid solutions; and (3) on glass surfaces incontact with water. DecaBDE dissolved in toluene was deposited on the solid substrate under a stream ofnitrogen (to evaporate the solvent) and then desiccated to remove any residual toluene. The adsorbed decaBDE on the solid substrate was then inundated with the aqueous test solution, followed by irradiation for the duration of the test period. In all experiments, natural sunlight (location, 40° 26’ N, 86° 55’ W) was used. The extent of degradation was determined using HPLC with UV detection or by GC/MS. In the first experiment, solar irradiation of decaBDE adsorbed onto humic acid-sand indicated that the photolysis of decaBDE was slow. After 96 hours of exposure to sunlight, 88% of initial decaBDEremained on the coated sand. There is some evidence that lower-brominated congeners (e.g., BDE 155)were formed in the experiment (EU 2002). In the second experiment, decaBDE was adsorbed on glasstubes containing a humic acid. In this study, the concentration of decaBDE decreased relatively quicklyover the first 24 hours of exposure, after which, the concentration remained stable. Bromide ion accumulated at an almost linear rate from start to end of the 72-hour exposure period. Approximately70% of the initial decaBDE remained after the 72-hour exposure. The difference in kinetics (for thedisappearance of decaBDE vs. the appearance of bromide ion) suggests that after the initial degradation ofdecaBDE, bromide ion was generated by the degradation of lower-brominated diphenyl ether congenerproducts (possibly octa- and nonaBDEs). Bromide ion mass balance for the system indicated that 70% ofthe total bromine present was accounted for by decaBDE or bromide, with the remaining 30% present asunidentified compounds. In the third experiment, Hua et al. (2003) investigated the photodegradation ofdecaBDE adsorbed on glass tubes, which were filled with aqueous solutions (without humic acid). Theresult of this test showed a much more rapid loss of decaBDE than found in the analogous test usinghumic acid solutions. Approximately 29% of the initial decaBDE present remained after 72 hours. Therate of decaBDE loss and bromide ion accumulation was relatively constant over the entire 72-hour testperiod. Mass balance indicated that approximately 50% of the total bromine was present as eitherdecaBDE or bromide ion, while the remaining 50% was possibly unidentified nona- and octaBDEcongeners. The difference between the tests using glass tubes with and without humic acid solution is possibly due to the absorption of light by humic acids, which may attenuate the degradation process. These studies indicate that adsorbed decaBDE may undergo photolysis forming octa- and nonaBDEsunder somewhat environmentally relevant conditions. Lower-brominated diphenyl ether congeners are PBDEs 3796. POTENTIAL FOR HUMAN EXPOSUREalso formed although only to a minor extent. These tests do not provide evidence that lower-brominated diphenyl ethers (e.g., tetra- and pentaBDEs) are a major degradation product of decaBDE (EU 2002). There is also insufficient information from these studies to estimate the rate of photolysis or ifintermediate degradation products build up after long-term exposures (EU 2002). Söderström et al. (2004) examined the time course of photolysis of decaBDE (BDE 209) in toluene, on silica gel, sand, sediment, and soil using artificial sunlight and on the natural matrices (e.g., sediment, soil, and sand) using natural sunlight. On natural samples, BDE 209 was first dissolved in toluene and then deposited on the natural matrix. The toluene was allowed to evaporate, and then the sample was reconstituted with water to resemble natural conditions. BDE 209 was photolytically labile and formed debromination products in all matrixes studied. Nona- to tetraBDEs were formed as well as some PBDFs. The half-lives in toluene and on silica gel were <15 minutes, and half-lives on other matrices ranged from 40 to 200 hours. No differences were observed in the debromination patterns under differentmatrices or light conditions. These experiments show that photolytic debromination of BDE 209 is apossible pathway for the formation of more bioavailable, lower-brominated PBDEs. However, the mostcommonly found BDEs in environmental samples (e.g., BDE 47, BDE 99, and BDE 100) were onlyformed to a minor degree (Söderström et al. 2004). Following the methodology described for decaBDE, photolysis experiments were conducted on BDE 47(EU 2002). BDE 47 was adsorbed on glass tubes filled with an aqueous solution and exposed to naturalsunlight. After 72-hours of exposure, 30% of the initial BDE 47 remained. The rate of disappearance ofBDE 47 was comparable to that found for decaBDE under similar test conditions. Accumulation ofbromide was initially slow with the rate increasing after 24 hours while the disappearance of BDE 47 wasinitial rapid over the first 24 hours. Using GC/MS, the authors concluded that 2,4,4’-triBDE was beingformed during this reaction and that removal of bromine atoms ortho to the ether functionality may be asignificant reaction pathway for removal of bromine atoms under the conditions of this study. This studysuggests that adsorbed BDE congeners, like decaBDE, may undergo photolysis under somewhatenvironmentally relevant conditions (EU 2002).PBDEs are not expected to undergo abiotic hydrolysis under environmental conditions due to the lack ofhydrolysable functional groups (Wolfe and Jeffers 2000).PBDEs are unlikely to biodegrade rapidly in the environment under aerobic conditions. PentaBDE did not undergo biodegradation (determined by CO2 evolution) after 29 days in an Organisation for Economic PBDEs 3806. POTENTIAL FOR HUMAN EXPOSURECo-operation and Development (OECD) 301B ready biodegradation test (EU 2001). The substance testedwas a composite sample from two producers with the following composition: 33.7% tetraBDE, 54.6%pentaBDE, and 11.7% hexaBDE. The test was extended to 93 days to allow sufficient opportunity foradaptation to occur. At the end of 93 days, 2.4% of the theoretical amount of CO2 had been evolved. Thus, pentaBDE was determined to be not readily biodegradable. No degradation (as oxygen uptake) was seen for octaBDE after 28 days in an OECD 301D ready biodegradation test (EU 2003a). Thus, octaBDEwas determined to be not readily biodegradable. The biodegradability of decaBDE has been studied under aerobic conditions using an activated sludge inoculum (EU 2002). DecaBDE at 100 mg/L was incubated with activated sludge (at 30 mg/L) over a 2-week period using a method similar to an OCED 301C MITI test. No degradation (as measured by biochemical oxygen demand) was observed. Thus, decaBDE was determined to be not readily biodegradable.No data on biodegradation of pentaBDE and octaBDE commercial mixtures under anaerobic conditionsare available. An anaerobic degradation study was carried out with BDE 47 using a mixture of14C-labeled and unlabeled compound (EU 2003a). The test was carried out using a sediment-water(Schuykill River, Pennsylvania) inoculum. After 32 weeks, it appeared that no significant degradation ofBDE 47 had occurred. However, the analytical method (i.e., HPLC using radiometric detection) used in this test indicated that some unidentified products had been formed in samples taken after 32 weeks.From these results, it is clear that BDE 47 has the potential to degrade slowly under anaerobic conditions(EU 2003a). Rayne et al. (2003b) reported that 4,4’-diBDE undergoes reductive debromination underanaerobic conditions. Debromination proceeds with replacement of a bromine atom by a hydrogen atom. The authors suggest that anaerobic debromination may sequentially debrominate BDE 15 to the parentdiphenyl ether. The anaerobic biodegradability of 14C-labeled decaBDE was studied over a period of 32 weeks (EU 2002). The test chambers consisted of 500 mL bottles containing 300 mL of sediment (Schuykill River, Pennsylvania) prepared under anaerobic conditions. The test chambers were incubated at 25°C and in the dark during the test. After the 32-week period, <1% of the total radioactivity added was found as 14CO2 and 14CH4 indicating that essentially no mineralization had occurred. GC/MS results showed no evidencefor the formation of lower-brominated congeners from decaBDE under the conditions of this test (EU2002). PBDEs 3816. POTENTIAL FOR HUMAN EXPOSURE6.3.2.3 Sediment and SoilInformation on the transformation and degradation of PBDEs in soil is limited. The extent to which PBDEs undergo direct photolysis in soils and sediment is unknown. However, sunlight would onlypenetrate the uppermost few millimeters of soil and will not impact sediment. Photolysis of PBDEs ispossibly important for land-applied sewage sludge contaminated with PBDEs. However, no information was available on this possibility. Based on studies in water, most PBDEs biodegrade slowly in soils orsediment under aerobic or anaerobic conditions. The anaerobic biodegradation of BDE 47, BDE 99, and BDE 209 was studied using microcosms prepared from loam sediment (pH 6.3, 16.4% organic carbon)obtained from a pond located in West Lafayette, Indiana (Tokarz et al. 2008). After an 8-month incubation period, microcosms containing BDE 47 showed variable losses (up to 30% of the initiallyapplied amount) of the parent congener without concurrent increases in expected debromination products, suggesting that other degradation mechanisms other than reductive debromination may have occurred. Only about 3% degradation of BDE 99 was observed after 8 months, with BDE 28 being the mostimportant debromination product. After 10 months, only slight decreases in the initial BDE 209 concentration was observed in six microcosms, and the half-life of this congener was estimated to range from 6 to 50 years; however, some aged microcosms exhibited greater degradation after 3.5 years, yielding nine degradation products (Tokarz et al. 2008). 6.3.2.4 Other MediaThe bacteria, Pseudomonas aeruginosa, that was isolated from an electronic waste dismantling area was capable of degrading BDE 209 under aerobic conditions, especially in the presence of co-metabolic substrates such as glucose (Shi et al. 2013). Nonabromodiphenyl ethers (BDE 208, BDE 207), fouroctabromodiphenyl ethers (BDE 203, BDE 202, BDE 197, BDE 196), and one heptabromodiphenylethers (BDE 183) were noted as degradation products.6.4 LEVELS MONITORED OR ESTIMATED IN THE ENVIRONMENTReliable evaluation of the potential for human exposure to PBDEs depends in part on the reliability ofsupporting analytical data from environmental samples and biological specimens. Concentrations ofPBDEs in unpolluted atmospheres and in pristine surface waters are often so low as to be near the limits of current analytical methods. In reviewing data on PBDEs levels monitored or estimated in theenvironment, it should also be noted that the amount of chemical identified analytically is not necessarily PBDEs 3826. POTENTIAL FOR HUMAN EXPOSUREequivalent to the amount that is bioavailable. The analytical methods available for monitoring PBDEs ina variety of environmental media are detailed in Chapter 7.Monitoring studies indicate that PBDEs are transported globally. Atmospheric, water, and biotaconcentrations of PBDEs tend to be dominated by lower-brominated congeners (e.g., BDE 47). Sediments tend to be dominated by higher-brominated congeners (e.g., BDE 209). Biota monitoringstudies indicate that PBDE concentrations have increased since the late 1970s, with lower-brominatedcongeners (e.g., BDE 47) being preferentially bioconcentrated. Studies indicate that PBDEconcentrations increase with respect to trophic level; organisms that reside higher on the food chain tend to have higher concentrations of PBDEs. 6.4.1 AirPBDEs will exist in both the vapor and particulate phase in both indoor and outdoor air (Harrad et al. 2004). The higher-brominated congeners (hepta-deca) have lower vapor pressures and partition more to the particulate phase, while the lower-brominated substances have a greater tendency to partition to the vapor phase. Concentrations of PBDEs in outdoor air in the United States are typically in the range of20–200 pg/m3, with BDE 47 and BDE 99 being the congeners most often detected (EPA 2010). Monitoring data from the 1990s showed infrequent detections of decaBDE; however, more recentmonitoring data have shown an increase in the frequency of detection of this substance both in outdoorand indoor air samples (EPA 2010; Hoh and Hites 2005). Monitoring data from Europe and Asia suggestlower concentrations of PBDE in air samples as compared to data obtained in the United States. Representative concentrations of PBDEs in outdoor air samples obtained at various locations aresummarized in Table 6-2. Air samples obtained from urban (Chicago, Illinois), rural (Sleeping BearDunes, Michigan and Sturgeon Point, New York), and remote (Eagle Harbor, Michigan) shorelines of theU.S. Great Lakes all contained quantifiable concentrations of BDE 47, BDE 99, BDE 100, BDE 153, and BDE 154 (Dodder et al. 2000; Strandberg et al. 2001). The most significant congeners were BDE 47 and BDE 99. Air measurements were averaged over a 3-year period between 1997 and 1999. Theconcentration of total PBDEs ranged from 5.5 pg/m3 in rural environments to 52 pg/m3 in urban air fromChicago, Illinois. The concentration of BDE 47 was 48 pg/m3 observed near Chicago, Illinois. Theaverage concentration of decaBDE at the remote and rural locations was <0.10 pg/m3 for each of the years investigated. The average concentration of decaBDE in the particulate phase at the urban location rangedfrom 0.20 to 0.35 pg/m3 (Strandberg et al. 2001). PBDEs 3836. POTENTIAL FOR HUMAN EXPOSURETable 6-2. Concentrations (pg/m3) of Several Polybrominated Diphenyl Ethers(PBDEs) in Air SamplesLocation BDE 47 BDE 99 BDE 100 BDE 209 ΣPBDEsa ReferenceUrban, United States48 25 3.0 No data 77* Dodder et al. 2000Rural, UnitedStates6.2–9.2 4.3–5.0 0.6–0.9 No data 2–4.8* Dodder et al. 2000Remote, United States3.7 2.6 0.33 No data 6.9* Dodder et al. 2000Alert, NorthwestTerritories CanadaNo data No data No data No data 1–28 Alaee et al. 1999Eagle Harbor,Wisconsin2.9 2.1 0.28 <0.10 5.5* Strandberg et al. 2001Chicago, Illinois 3.9–42 2.4–15 0.68–3.3 1.5–878 13–980 Hoh and Hites 2005Sleeping BearDunes, Michigan0.51–27 0.32–23 0.030–5.1 <0.29–21 1.4–61* Hoh and Hites 2005Bloomington,Indiana1.9–21 1.2–11 0.26–2.75 <0.29–21 6.4–44* Hoh and Hites 2005Rohwer, Arkansas 1.2–42 0.87–35 0.17–3.7 <0.10–135 2.7–165* Hoh and Hites 2005Cocodrie,Louisiana2.0–24 0.89–11 0.21–2.7 <0.10–14 5.1–42* Hoh and Hites 2005Sturgeon Point,New York3.8 2.8 0.39 <0.10 7.2* Strandberg et al. 2001Sleeping BearDunes, Michigan8.4 5.3 0.80 <0.10 15* Strandberg et al. 2001Chicago, Illinois 33 16 2.0 0.20–0.35 52* Strandberg et al. 2001Ammarnäs,Sweden6.3 1.6 0.4 No data 8.3 de Wit 2000, 2002Hoburgen,Sweden0.7 0.35 0.07 No data 1.1 de Wit 2000, 2002Stoke Ferry,United Kingdom4.7–50 5.5–13 1.1–3.9 No data 6.7–58 Peters et al. 1999Hazelrigg, United Kingdom3.2–61 3.1–22 0.62–5.4 No data 4.1–69 Peters et al. 1999Dunai Island,RussiaNo data No data No data No data 1–8* Alaee et al. 1999Arctic 2.2–2.8 2.0–2.3 0.40–0.47 1.0–1.8 6.7–8.6* Hung et al. 2010aΣPBDEs is the sum of BDE 47, BDE 99, and BDE 100, but if more congeners are included, this is marked with an asterisk (*).BDE = brominated diphenyl ether PBDEs 3846. POTENTIAL FOR HUMAN EXPOSUREPBDEs were monitored at five different locations (Chicago, Illinois; Sleeping Bear Dunes, Michigan;Bloomington, Indiana; Rohwer, Arkansas; and Cocodrie, Louisiana) across the United States from 2002 to 2003 (Hoh and Hites 2005). Total PBDE concentrations at the urban site (Chicago, Illinois) were 3– 6 times greater than the other locations, with BDE 47, BDE 99, BDE 100, and BDE 209 being the mostabundant congeners at all five locations. DecaBDE concentrations as high as 960 and 410 pg/m3 were observed in Chicago on two dates in 2003.Throughout the year of 1997, air samples were taken from a rural site in southwestern England calledStokes Ferry and a semirural site in northwestern England called Hazelrigg and analyzed for PBDEs (Peters et al. 1999). Tri- and heptaBDEs were detected; the combined concentrations of BDE 47,BDE 99, and BDE 100 ranged from 7 to 69 pg/m3 at Hazelrigg and from 6 to 58 pg/m3 at Stoke Ferry (de Wit 2002). PBDEs have also been measured in air samples taken from remote stations in the Arctic (e.g.,Alert, Northwest Territories, Canada; Dunai Island, eastern Siberia, Russia) between January 1994 and January 1995 (de Wit 2002). The total concentration of several di- to hexaBDEs ranged from 1 to 4 pg/m3 at Alert for the majority of the year; however, in July 1994, the concentration was 28 pg/m3. AtDunai, the major congeners found were BDE 47 and BDE 99. In Sweden during 1990–1991, air samplescollected from Ammarnäs in the northern mountains and Hoburgen on the southern tip of Gotland in theBaltic Sea, had measurable amounts of BDE 47, BDE 99, and BDE 100 (de Wit 2002). Total PBDEconcentrations were approximately 1 and 8 pg/m3, respectively. The concentration of BDE 47 was found to be highest in the gas phase, while BDE 99 and BDE 100 were highest in the particulate phase. No decaBDE was found, although the limit of detection limit for decaBDE is much higher than for the lowerbrominated diphenyl ethers.Indoor air concentrations of PBDEs vary depending upon potential sources such electronics or foams usedin upholstery stuffing for furniture that were treated with PBDEs (EPA 2010; Harrad et al. 2006; Hazratiand Harrad 2006). Harrad et al. (2004) found a significant positive correlation between PBDEconcentrations in indoor air and both the number of electrical appliances and the number of chairscontaining FPUF. Concentrations of tetra- and pentabrominated congeners (BDE 47, BDE 99, and BDE100) in indoor air were always higher than those detected in outdoor air. On average, indoor airconcentrations were 150, 120, and 140 times higher than outdoor air for BDE 47, BDE 99 and BDE 100, respectively. Indoor air concentrations of PBDEs also tended to be higher in workplace environments as compared to domestic residences; however, these concentrations could vary substantially from one roomto another (Harrad et al. 2004). Air samples collected from 31 homes, 33 offices, and 25 automobiles in the West Midlands, United Kingdom were analyzed for the presence of PBDEs (Harrad et al. 2006). PBDEs 3856. POTENTIAL FOR HUMAN EXPOSURETotal PBDE concentrations in samples obtained from homes ranged from 4 to 245 pg/m3 (24 pg/m3, median), while concentrations in offices and cars ranged from 10 to 1,416 (71 pg/m3, median) and from4 to 1,416 pg/m3 (41 pg/m3, median), respectively. PBDE congeners 47 and 99 were reported as themajor contributors to the overall total (Harrad et al. 2006). Although the previous study reported astatistically significant positive correlation between the PBDE concentrations and the number of electricaldevices and FPUF-containing chairs (Harrad et al. 2004), no clear statistical evidence of such correlations were reported in these indoor air environments (Harrad et al. 2006; Hazrati and Harrad 2006). PBDEconcentrations in the air of one office did drop dramatically after an older computer was replaced by a relatively newer one and a statistically significant positive correlation was observed between automobile age and PBDE concentrations when two highly contaminated outliers were removed from the analysis. Further observations indicated a seasonal variation of the PBDE concentrations in indoor air, with higherconcentrations being observed during the summer months (Hazrati and Harrad 2006). BDE 47, BDE 99, BDE 100, BDE 183, and BDE 209 were detected in 63, 22, 29, 32, and 42% of air samples taken aboard aircraft during routine flights (Allen et al. 2013). Concentrations ranged from below the detection limitsto a maximum concentration of 2,100,000 pg/m3 for BDE 209.Concentrations of PBDEs were measured in floor dust, indoor air, ventilation filter dust and carpets in 10 buildings located in Michigan (Batterman et al. 2010). Median concentrations of total PBDEs werereported as 8,754 ng/g in settled dust, 1,250 pg/m3 vapor-phase air, and 155 pg/m3 particulate-phase air.The highest concentrations of PBDEs were generally noted in rooms that contained sources such ascomputer servers. Monitoring data from one building that was built in 2006 indicated a very lowconcentration of PBDEs in settled dust at the time it was constructed (145 ng/g); however, these concentrations increased exponentially to >10,000 ng/g 5–8 months after the building was opened. Despite the voluntary phase-out of pentaBDE and octaBDE in 2004, high concentrations of congenersassociated with these mixtures (BDE 47, BDE 99, BDE 100, BDE 153, and BDE 203) were detected in dust and air samples, suggesting that there are significant sources of these PBDEs in products that remainin the market and that were used in this building even after the phase out in 2004.6.4.2 WaterDue to the hydrophobic nature of PBDEs, this class of compounds is expected to be present in water atvery low concentrations or at concentrations below the limit of detection of acceptable analyticalmethods. In 1999, the concentration of PBDEs in Lake Ontario surface waters ranged between 4 and 13 pg/L with ~90% in the dissolved phase (Luckey et al. 2001). BDE 47 and BDE 99 were the most PBDEs 3866. POTENTIAL FOR HUMAN EXPOSUREabundant congeners, together making up >70% of the total PBDEs. Streets et al. (2006) reported 18 and 3.1 pg/L average concentration of ƩPBDEs (BDE 47, BDE 66, BDE 100, BDE 99, BDE 85, BDE 154, and BDE 153 of pg/L) in Lake Michigan Water for the dissolved phase and particulate phase, respectively. In Japan, PBDEs were not detected in 75–200 water samples (ENVIRON 2003a). In nineEnglish freshwater lakes grab samples were obtained from April 2008 to February 2012, averageconcentrations of the sum total of PBDEs, which included BDE 17, BDE 28, BDE 49, BDE 47, BDE 66, BDE 100, BDE 99, BDE 85, BDE 154, and BDE 153, ranged from 41.5 to 73.3 pg/L (Yang et al. 2014). The EPA utilized data from the San Francisco Monitoring Program for Trace Substances in surface waterto estimate possible concentrations of PBDEs that might be found in drinking water. Thirty-three watersamples were obtained from the San Francisco Estuary, with total PBDEs concentrations ranging from3 to 513 pg/L, and a mean concentration of 146.2 pg/L (Oros et al. 2005). It was reported that the mostfrequently detected congeners were BDE 47, BDE 99, and BDE 209. The source of these PBDEs was most likely effluent from municipal treatment plants. Concentrations of PBDEs monitored in theSpokane River in the state of Washington exhibited seasonal variation (Furl and Meredith 2010). Dissolved PBDE concentrations collected in the fall of 2005 were approximately 6 times greater(926 pg/L) as compared to concentrations observed in the spring of 2006 (146 pg/L). The variation in PBDE concentrations were likely a result of dilution of local sources in the spring from increased flowdue to snowmelt in the upper watershed. 6.4.3 Sediment and SoilHale et al. (2002) reported the concentration of PBDEs in soil samples collected in the vicinity of apolyurethane foam-manufacturing facility. Concentrations in these soils are likely to be higher than those to be expected in rural and potentially urban areas of the United States. Total PBDE concentrations in these samples ranged from not detected to 76 μg/kg dry weight. BDE 99 was the predominant congenerin soil followed by BDE 47 and BDE 100. Concentrations of total PBDEs in soil obtained from a largeautomotive shredding and metal recycling facility in Brisbane, Australia ranged from 29 to 726 ng/g dryweight as compared to background concentrations at an uncontaminated site (0.2–2 ng/g dry weight)(Hearn et al. 2013). BDE 209 was the predominant congener in dust, soil, and air at the facility.In Eastern China, near dismantling areas for waste electrical and electronic equipment that are considereda potential exposure source of PBDEs, 45 farmland soil samples were collected from 12 locations (Donget al. 2014). The farmland soil PBDE concentrations ranged from 2.96 to 200 ng/g and air concentrations PBDEs 3876. POTENTIAL FOR HUMAN EXPOSUREranged from 884 to 2,791 pg/m3. The PBDE concentrations in high-mountain pasture soil, grass, and milk from grazing cows in the Italian Alps were analyzed (Parolini et al. 2012). Thirteen BDE congeners were investigated, including BDE 17, BDE 28, BDE 71, BDE 47, BDE 66, BDE 100, BDE 99, BDE 85, BDE 154, BDE 153, BDE 138, BDE 183, and BDE 190. Average ƩPBDE concentrations in soil from0 to 7 cm soil depth were 0.43–1.55 ng/g dry weight, 1.7–8.2 ng/g dry weight in grass vegetation, and 0.659–1.576 ng/g dry weight in milk. The average total PBDE concentration in soil samples collected from the Ny-Alesund region of the Arctic was reported as 0.042 ng/g dry weight (Wang et al. 2015). BDE 99 was the predominant congener, with an average value of 0.0097 ng/g dry weight.The EPA used 33 surface soil measurements taken from 15 states in the United States to estimate ingestion rates and dermal exposure rates of PBDEs from soil. The concentration of 30 total BDEs in thesoils was reported to average 103 ng/g dry weight, with a geometric mean (GM) concentration of 5.3 ng/g(EPA 2010). BDE 47 (1.9 ng/g), BDE 99 (3.6), BDE 100 (0.4), BDE 153 (5.7), BDE 154 (4.8), BDE 183 (37.4), and BDE 209 (15.3) were included in this evaluation.Sediment concentrations of PBDEs tend to be dominated by higher-brominated congeners (e.g., BDE 209) (deWit 2002). Temporal trends suggest that concentrations of PBDEs in sediments areincreasing. Burdens of PBDEs in sediment appear to be a function of distance from the source and theirorganic carbon content (Hale et al. 2003). Representative concentrations of PBDEs in sediment samples are summarized in Table 6-3. A 20-year field study regarding the land application of Class B biosolids to a site located in Arizona was discussed by Quanrud et al. (2011). Risk assessments were made based on the intake of compounds viainhalation, dermal sorption, or ingestion. PBDE concentrations were detected, primarily in the surface30-cm depth sample, and surface accumulation of PBDEs occured due to their hydrophobic nature, which resulted in sorption to colloids. The maximum amount of PBDE detected was 80 ng/g soil as congenerBDE 209. A risk evaluation of PBDEs based on Hazard Indices indicated that the health risk to humansof PBDEs was negligible when all three routes of exposure were considered. Li et al. (2006) collected 199 sediment samples from 16 locations of Lake Superior, Lake Michigan, Lake Huron, Lake Erie, and Lake Ontario and analyzed these samples for PBDE concentrations. Theconcentrations of PBDEs in surface sediment ranged from 0.5 to 6.7 ng/g dry weight tri- to heptaBDEcongeners and from <4 to >240 ng/g dry weight for BDE 209. PBDEs were detected in 100% ofsediment samples obtained from the San Francisco Bay with total PBDE concentrations ranging from PBDEs 3886. POTENTIAL FOR HUMAN EXPOSURETable 6-3. Concentrations (ng/g) of Several Polybrominated Diphenyl Ethers(PBDEs) in Sediment and Suspended Particulate SamplesSample type Location BDE 47 BDE 99 BDE 100 ΣPBDEs BDE 209 ReferenceSediment Lake Superior No data No data No data No data 12 Zhu and Hites2005Sediment Lake Michigan No data No data No data 2.6a 320 Zhu and Hites2005Sediment Lake Michigan No data No data No data 1.67–3.97a 43.9–95.6 Song et al. 2005Sediment Lake Huron No data No data No data 1.02–1.87a 21.5–36.0 Song et al. 2005Sediment Lake Erie No data No data No data 1.1a 40 Zhu and Hites2005Sediment Lake Ontario No data No data No data 2.8a 14 Qui et al. 2007Sediment Hadley Lake,Indiana16±2 (dw) 37±8 (dw)7.1±1.5(dw)584 (dw)b 480±170 (dw)Dodder et al.2002Sediment Baltic Sea ND–3.4 ND–2.4 ND–1.3 ND–5.4 ND Nylund et al.1992Sediment Upstreamplastics plant,Sweden3.7 8.8 1.6 14.1 No data Sellström and Jansson 1995Sediment Downstreamplastics plant,Sweden780 1,200 270 2,250 No data Sellström and Jansson 1995Sediment River Viskan (Sweden), upstream and <2–50 <1–53 <0.4–19 ND–120 ND–16,000 Sellström et al.1998adownstreamtextile industriesSediment 22 European river mouths<0.17–6.2 (dw)<0.19– 7.0 (dw)No data No data <0.51– 1,800de Wit 2002Sediment Seven rivers,Great Britain<0.3–368 (dw)<0.6– 898 (dw)No data No data <0.6–3,190 Allchin et al.1999Sediment Netherlands,several sites0.3–7.1 (dw)<0.2–9 (dw)No data No data <4–510(dw)de Boer et al.2000bSuspended particulatesNetherlands,several sites<2–9 (dw) <0.1–23 (dw)No data No data <9–4,600 (dw)de Boer et al.2000baTri- to hepta-PBDE congeners.bIncludes sum of BDE 47, BDE 99, BDE 100, BDE 209, and other congeners (not specified).BDE = brominated diphenyl ether; dw = dry weight; ND = not detected PBDEs 3896. POTENTIAL FOR HUMAN EXPOSURE2.1 to 8.0 ng/g dry weight and a median concentration of 4.3 ng/g dry weight (Klosterhaus et al. 2012). BDE 209 was the primary congener of each sample, accounting for approximately 38–68% of the totalPBDE amount. In the United States, Dodder et al. (2002) analyzed four surficial sediments from Hadley Lake (Indiana). This lake is in the vicinity of a production point source. DecaBDE (BDE 209) was the major congenerdetected at concentration ranging from 19 to 36 μg/kg (ng/g) dry weight. Other congeners detected (in decreasing order: BDE 99, BDE 153, BDE 154, BDE 47, and BDE 100) were <5 μg/kg (ng/g) dry weight. PBDEs were above the detection limit (i.e., 0.5 μg/kg [ng/g] dry weight) in 22% of surficial sedimentsamples (from 133 sites) in freshwater tributaries of Virginia (Hale et al. 2001b). BDE 47 was the predominant congener followed by BDE 99 and BDE 100. The maximum concentration detected in sediment was 52.3 μg/kg (ng/g) dry weight. Hale et al. (2002) reported that stream sediment adjacent to a former polyurethane foam production facility in North Carolina contained up to 132 μg/kg (ng/g) dry weight of pentaBDE. Since the phase out of PBDE-containing flame retardants, levels of these substances have begun declining in sediment and other environmental media. Sutton et al. (2015)reported that levels of BDE 47 have declined by over one-third in sediment samples obtained from theSan Francisco Bay from 2000 to 2012; however, no decline of BDE 209 was evident, presumably since decaBDE was not phased out until 2013. In Japan, tetra-, penta-, hexa-, and decaBDEs have been detected in river sediments (Watanabe et al.1986, 1987, 1995). The combined concentrations of tetra- and pentaBDEs ranged from 21 to 59 ng/g (μg/kg) dry weight. The concentration of decaBDE (BDE 209) ranged from <25 to 11,600 ng/g (μg/kg)dry weight (deWit 2002). In 1999, sediment samples from several locations in the Netherlands contained BDE 47, BDE 99, and BDE 209 (de Boer et al. 2000b). Concentrations ranged from 0.3 to 7.1 ng/g (μg/kg) dry weight for BDE 47, not detected to 5.5 ng/g (μg/kg) dry weight for BDE 99, and not detected to 510 ng/g (μg/kg) dry weight for BDE 209. The concentration of PBDEs in suspended particulatematter ranged from not detected to 9 ng/g (μg/kg) dry weight for BDE 47, not detected to 23 ng/g (μg/kg)dry weight for BDE 99, and not detected to 4,600 ng/g (μg/kg) dry weight for BDE 209 (de Boer et al. 2000b). The concentration of several brominated flame retardants was measured in sediments collectedfrom the mouths of major European rivers (de Wit 2002). Elevated concentrations of BDE 47 and BDE 99 were found in Humber and Mersey rivers (Great Britain). In two rivers of the Netherlands, the sum of BDE 47 and BDE 99 ranged from 1.61 to 13.1 ng/g (μg/kg) dry weight. The highest hexaBDEconcentrations (as BDE 153) were found in the river Seine (France), three rivers in the Netherlands, and the rivers Schelde (Belgium), Forth (Great Britain) and Ems (Germany); the concentration of PBDEs 3906. POTENTIAL FOR HUMAN EXPOSUREBDE 153 ranged from 0.013 to 0.056 ng/g (μg/kg) dry weight in these sediments. The concentrations ofdecaBDE were highest in sediment from the Seine, ranging from 2.4 to 3.9 ng/g (μg/kg) dry weight. Theconcentrations of decaBDE in River Mersey (Great Britain), Schelde, and River Liffey (Ireland) rangedfrom 34 to 1,800 ng/g (μg/kg) dry weight. In the southern Baltic Sea (Bornholm Deep), the upper layer ofsediment was analyzed for BDE 47, BDE 99, and BDE 100; the combined concentration of these three congeners was 0.52 ng/g (μg/kg) dry weight (Nylund et al. 1992). A well-studied sediment core collected from the southern part of the Baltic Sea proper was analyzed forPBDEs and a number of organochlorine contaminants (Nylund et al. 1992). The retrospective temporaltrend from 1939 to 1987 showed that the PBDE concentrations (i.e., sum of BDE 47, BDE 99, and BDE 100) have increased with a sharp increase after 1980. The PBDE concentration in the sample from1989 was 2.9 ng/g (Nylund et al. 1992). Measurable amounts of BDE 28, BDE 47, BDE 66, BDE 99, and BDE 100 were found in sediment cores from a freshwater lake in Germany, the Wadden Sea (the Netherlands), and Drammenfjord (Oslo Fjord, Norway) (Zegers et al. 2000). Samples from the Drammenfjord and freshwater lake also contained BDE 153 and BDE 154, and the Wadden Sea and freshwater lake samples contained BDE 209. The lower-brominated PBDEs appear in the 1960s, and BDE 209 appears about 10 years later. The Drammenfjord sediment core shows increasingconcentrations of BDE 47 starting in the 1940s (range, 0.02–0.18 ng/g dry weight) and increasing concentrations of BDE 99 (range, 0.5–0.28 ng/g dry weight), BDE 100 (range, not detected–0.07 ng/g dry weight), and BDE 154 (range, not detected–0.06 ng/g dry weight) beginning in the 1950s up to 1999. In the sediment core from Lake Woserin, lower-brominated PBDE congeners were detected in the sedimenthorizons beginning in the late 1950s, and the concentrations increased until the late 1970s, and then leveled off when residues of BDE 209 first appeared. A similar leveling-off trend is also observed in theWadden Sea core (Zegers et al. 2000). It is important to note that this study identified the presence ofPBDEs compounds in sediments from the late 1950s and early 1960s. This is nearly a decade prior to anysignificant commercial production of these substances. The existence of PBDEs at these early dates maybe a result of vertical mixing of sediment cores or blurring of core horizons through burrowing activity ofbenthic organisms, or may lend some credibility to the likelihood that either the substances identified inthe environment as PBDEs are not necessarily PBDEs.6.4.4 Other Environmental Media Dust. The predominant PBDE exposure pathway for the general population of the United States is fromindoor dust (EPA 2010; Lorber 2008). Total PBDE (sum of BDE 28, 47, 66, 85, 99, 100, 153, 154, 183, PBDEs 3916. POTENTIAL FOR HUMAN EXPOSURE206, and 209) levels in dust samples obtained from 20 residences located near the Columbia River in Washington state ranged from 311 to 19,700 ng/g, with BDE 209 being the predominant congener(Shreder and LaGuardia 2014). Total PBDE concentrations in settled dust from 10 office buildingslocated in Michigan ranged from 1,340 to 38,900 ng/g (median, 15,800 ng/g) (Batterman et al. 2010). BDE 47, BDE 99, and BDE 209 had the highest concentrations for the individual congeners, with amaximum value of 29,000 ng/g measured for BDE 209 in one of the office buildings. A study of eightoffice buildings in Boston, Massachusetts had even greater PBDE concentrations in settled dust. BDE209 was detected in 100% of the samples at concentrations ranging from 912 to 106,204 ng/g (Watkins etal. 2011). Total pentaBDE congeners ranged from 141 to 61,264 ng/g in these office buildings. Lowerconcentrations of the lesser brominated congeners, Σtri–hexaBDE, were detected in dust samples in Europe as compared to the United States and Canada (Harrad et al. 2008). Concentrations of BDE 209 in dust samples are similar in the United States and the United Kingdom. Congener pattern analysis ofindoor dust suggests that North American dusts are contaminated with decaBDE and pentaBDEcommercial formulations, whereas U.K. dusts are predominantly contaminated with decaBDE. Forexample, the average concentration of total PBDE congeners in eight homes located in West Midlands,United Kingdom was about 215 ng/g, with a range of 16.2–625.4 ng/g (Harrad et al. 2006). Dust samplesobtained from Spain and Belgium had total PBDE ranges of 2.9–380.2 and 6.2–384.8 ng/g, respectively(Harrad et al. 2006). Concentrations in North America also seem to be greater than even source-dominated areas in other parts of the world. The mean total concentration of 10 PBDE congeners in dustsamples obtained from an office building located near a large automotive shredding and metal recyclingfacility in Brisbane, Australia was 2,014 ng/g (Hearn et al. 2013). Food. Food ingestion typically accounts for <20% of the total PBDE intake for adults in North America(EPA 2010; Lorber 2008); however, it accounts for the majority of intake for the European population, with the exception of BDE 209 where dust exposure is the primary source (Abdallah and Harrad 2014;Trudel et al. 2011). Schecter et al. (2006) measured concentrations of 13 PBDE congeners in 62 food samples in the United States. Concentrations of total PBDEs ranged from 7.9 pg/g in milk to 3,762 pg/g in canned sardines. Fish, meat, and dairy products tended to have the highest concentrations of PBDEs. The results of these measurements for meat and fish are reproduced in Table 6-4. A survey of threecategories of baby food (formula, cereal, and puree) from the United States and China found the median concentrations of total PBDE (sum of BDE 17, BDE 28, BDE 47, BDE 49, BDE 100, BDE 153, BDE183, and BDE 209) were 21 and 36 pg/g in foods purchased in the United States and China, respectively(Liu et al. 2014). PBDEs 3926. POTENTIAL FOR HUMAN EXPOSURETable 6-4. PBDE Concentrations (pg/g Wet Weight) in 18 U.S. Meat and Fish SamplesaLipid BDE BDE BDE BDE BDE BDE BDE BDE BDE BDE BDE BDE BDE TotalSample (%) 17 28 47 66 77 85 99 100 138 153 154 183 209 PBDEsb U.S. meat samplesBacon A 52.3 ND ND ND ND ND ND ND ND ND ND ND ND ND 165(5.2) (7.1) (78.8) (5.2) (5.2) (5.2) (28.8) (6.8) (5.2) (5.2) (5.2) (5.2) (166.6)Bacon B 43.4 ND ND ND ND ND NA ND ND ND ND ND ND ND 39(0.4) (2.1) (19.9) (0.4) (0.2) (15.6) (2.8) (0.4) (1.1) (0.9) (1.7) (32.8)Bacon C 35.3 0.7 ND 30.1 NA NA 1.4 16.8 4.8 ND 4.5 2.8 14.3 28.4 105(2.0) (0.7)Beef 30.7 ND 59.7 87.5 ND ND ND 35.5 6.2 ND 6.8 4.6 ND ND 258(ground) A (3.1) (3.1) (3.1) (3.1) (3.1) (4.2) (95.7)Beef 13.6 0.2 ND 23.4 0.5 NA NA 32.3 4.5 0.4 4.7 2.5 NA 9.7 79(ground) B (0.7)Beef 13.7 ND ND 35.1 ND NA 1.7 40.3 6.9 ND 4.9 3.7 3.8 ND 105tenderloin (1.4) (1.5) (1.4) (1.4) (11.1)Chicken 4.9 ND 0.5 60.5 NA NA NA 128 17.1 2.2 12.0 10.8 3.2 48.5 283breast (0.04)Duck 75.1 ND ND 286 2.7 ND 15.2 609 122 7.3 52.3 42.9 31.6 113 1,283(0.5) (3.0) (0.3)Ground 7.3 ND ND 11.0 ND NA ND 18.9 4.6 ND 4.1 2.6 5.8 80 129chicken (0.7) (1.5) (0.7) (0.7) (0.7)Ground 19.7 ND ND ND ND ND 3.2 56.8 16.8 ND 9.6 6.3 ND ND 186lamb (2.0) (2.1) (23.0) (2.0) (2.0) (2.0) (2.0) (150.6)Ground 21.5 ND ND 53.8 ND NA 3.1 74.2 12.9 4.3 18.7 15.0 19.9 ND 221pork (2.2) (3.5) (2.2) (31.3)Ground 11.1 0.2 ND 98 0.8 ND NA 217 54.4 3.9 32.9 24.1 36.8 245 713turkey (0.5) (0.1)Pork 8.9 0.1 ND 6.9 NA NA NA 16.3 1.8 0.2 1.0 1.2 1.3 11.7 41(0.5)Pork 23.7 ND ND 387 ND ND 16.8 688 74.5 5.6 81.6 55.3 14.6 49.7 1,378sausage A (1.3) (6.9) (1.0) (0.3)Pork 24.4 ND ND 39.4 ND ND 2.6 71.6 8.3 ND 22.0 13.7 10.7 ND 244sausage B (2.4) (3.4) (2.4) (2.4) (2.4) (139)Sausage 26.2 ND ND ND ND NA 3.1 40.1 6.4 ND 5.9 4.9 6.9 ND 1,426A (2.6) (5.5) (34.8) (2.6) (2.6) (51.0)Sausage 28.5 ND ND 94.1 ND ND ND 43.7 8.3 ND 8.5 9.2 ND ND 195B (2.9) (3.2) (3.5) (2.9) (2.9) (2.9) (2.9) (41.7)Wieners 32.9 ND ND 386 1.4 ND 11.1 703 53.9 7.2 106 49.8 14.3 ND 1,348(0.3) (1.5) (0.2) (28.7)Mean 26.3 0.76 4.59 93.2 1.19 0.83 4.93 157 22.7 2.33 21.1 14 10.1 53.3 383Median 24.1 0.66 1.03 39.4 1.08 0.57 2.62 42 7.57 1.37 7.68 5.63 5.83 38.1 190Minimum 4.87 0.02 0.24 6.93 0.21 0.06 0.36 7.79 1.39 0.16 0.53 0.44 0.86 5.53 39Maximum 75.1 2.62 59.7 387 2.74 2.62 16.8 703 121 7.28 106 55.3 36.8 245 1,426U.S. fish samplesCanned 0.3 0.1 0.6 5.1 0.2 NA 0.2 3.2 0.6 ND 0.3 0.2 1.1 4.9 16.6tuna A (0.0)Canned 0.5 ND 0.2 2.1 0.2 NA ND 1.1 0.4 ND 0.2 0.3 2.1 8.8 15.5tuna B (0.1) (0.1) (0.1)Catfish A 11.1 4.6 6.4 372 4.3 NA NA 589 116 5.1 37.1 39.6 7.3 1269 2,450Catfish B 5.3 4.6 5.1 438 13.5 ND 41.6 834 102 7.9 49.9 45.8 4.9 ND 1,547(0.1) (15.9)Catfish C 5.2 2.2 3.7 137 0.7 ND 11.7 184 39.5 ND 15.8 15.2 ND ND 437(0.5) (2.7) (1.6) (49.4) PBDEs 3936. POTENTIAL FOR HUMAN EXPOSURETable 6-4. PBDE Concentrations (pg/g Wet Weight) in 18 U.S. Meat and Fish SamplesaSampleLipid (%)BDE 17BDE 28BDE 47BDE 66BDE 77BDE 85BDE 99BDE 100BDE 138BDE 153BDE 154BDE 183BDE 209TotalPBDEsb Catfish 5.7 1.1 3.7 197 6.3 NA 16.4 282 53.0 ND 18.4 21.3 3.8 22.7 627fillet (farm) (4.1)Halibut 0.2 0.6 4.1 76.6 2.8 NA ND 10.6 12.4 ND 1.1 2.6 1.8 11.4 124(0.1) (0.1)Herring 9.1 4.1 56.3 2,072 69.4 3.6 ND 267 221 ND 29.3 69.9 2.5 ND 2,809(0.9) (0.9) (26.4)Mahi 0.5 0.6 ND 24.1 2.0 NA 0.6 13.0 5.1 ND 1.4 4.9 4.3 ND 66(2.0) (0.8) (16.6)Salmon A 8.0 79.2 92.6 1,222 30.6 ND NA 93.2 348 ND 27.7 98.8 1.4 ND 1,999(0.2) (0.2) (9.0)Salmon B 13.9 118 142 2,081 59.1 ND NA 147 353 ND 36.6 142 ND ND 3,082(0.1) (0.2) (1.2) (7.0)Salmon C 10.3 18.4 49.4 1,103 35.3 ND ND 239 217 ND 18.3 45.1 ND ND 1,732(0.1) (0.1) (0.1) (1.3) (11.2)Salmon D 6.3 1.4 5.2 94.7 5.2 ND ND 15.4 7.1 ND 1.4 5.0 ND ND 141(0.9) (0.6) (0.6) (0.8) (9.1)Salmon E 12.3 1.7 20.4 356 ND ND ND 84.4 84.2 ND 10.1 29.8 ND ND 605(2.1) (1.2) (1.2) (1.2) (1.4) (29.2)Salmon 7.4 11.1 50.5 1,000 63.1 NA 7.9 410 210 ND 37.4 104 3.7 20.5 1,919fillet (farm) (1.4)ASalmon 6.9 2.3 27.9 517 24.3 NA ND 168 115 ND 16.0 35.8 1.7 681 1,590fillet (farm) B(0.7) (0.7)Sardines 9.6 3.3 53.6 2,748 85.6 ND ND 358 257 ND 51.9 139 ND ND 3,726(5.0) (1.0) (1.0) (3.2) (51.4)Shark 0.4 1.1 29.8 784 29.5 0.3 NA 57.8 608 0.4 112 291 2.0 5.4 1,920Shrimp 0.6 0.3 3.6 75.6 NA NA NA 9.4 14.3 ND 1.2 2.6 0.2 ND 108(0.1) (1.3)Tilapia 1.0 ND (0.1)ND (0.7)5.9 NA NA 0.1 1.3 0.6 ND (0.1)0.2 0.5 ND (0.2)ND (4.0)11Trout A 4.2 4.8 22.2 320 NA NA ND 79.8 66.5 0.2 11.8 26.3 4.4 ND 549(0.2) (26.7)Trout B 10.1 4.3 49.3 826 ND ND ND 128 198 ND 24.7 61.3 2.5 ND 1,319(5.6) (1.0) (1.0) (1.0) (42.9)Tuna 0.2 ND ND 16.6 0.7 NA ND ND 2.9 ND ND ND 0.5 23.4 48(0.1) (1.0) (0.0) (4.6) (0.1) (0.4) (1.0)Wild perch 1.2 ND (0.1)0.7 10.2 0.4 NA ND (0.1)2.3 2.1 ND (0.1)0.7 2.4 0.6 5.9 25Mean 5.43 11.01 26.19 603 20.8 0.78 4.29 166 126 0.89 21 49.3 2.08 91.8 1,120Median 5.52 1.97 5.77 338 5.23 0.30 0.35 88.8 75.3 0.33 15.9 28. 1.68 10.1 616Minimum 0.15 0.03 0.20 2.11 0.18 0.06 0.02 1.15 0.43 0.02 0.21 0.21 0.12 0.63 11.14Maximum 13.9 118 142 2,748 85.6 3.60 41.6 834 608 7.94 112 291 7.32 1,269 3,726aLimits of detection (LODs) are shown in parentheses. Total PBDE concentrations and statistics for each congener were calculated byassuming that nondetected concentrations were one-half the LOD; for calculations, these were treated as zero.bTotals rounded to the nearest whole number for hundreds and to the nearest decimal place for tens.NA = not available; ND = not detectedSource: Adapted from Schecter et al. (2006) PBDEs 3946. POTENTIAL FOR HUMAN EXPOSUREHuwe et al. (2002a) reported total PBDE concentrations in farm chickens raised in two different regionsof the United States. The total PBDE concentration of discrete samples of chickens raised in Arkansas was 39.4 ng/g whole weight, while one composite sample of chickens raised in North Dakota was 1.7 ng/g whole weight. In the United States, chickens fed ball clay (a sedimentary, kaolinite clay) andchickens bought in the grocery store were analyzed for total PBDEs (ENVIRON 2003b). BDE 99 was the dominant congener in all samples. Total PBDEs ranged between 4 and 35 ng/g lipid weight in chickens fed ball clay and 0.5 ng/g lipid weight in store-bought chicken. Ohta et al. (2002) determined the concentration of total PBDEs in vegetables and meat samples from Japan. The concentrations ofPBDEs in spinach, potato, and carrot were 134, 47.6, and 38.4 pg/g fresh weight, respectively. Thehighest concentrations of total PBDEs and BDE 47 were found in spinach. Interestingly, differentcongener patterns were found among the vegetables analyzed. Compared to root vegetables, which hadhigh concentrations of BDE 153, spinach (representing a leafy vegetable) might be strongly influenced byPBDE contamination in air. The concentrations of PBDEs in pork, beef, and chicken were 63.6, 16.2, and 6.25 pg/g fresh weight, respectively. PBDE concentrations were highest in pork samples; however, the reason for this is unknown (Ohta et al. 2002). Bocio et al. (2003) determined the concentrations ofPBDEs in food samples from Catalonia, Spain during 2000. The highest concentration of total PBDEswas found in oils and fats (587.7–569.3 pg/g), followed by fish and shellfish (333.9–325.3 pg/g), meatand meat products (109.2–102.4 pg/g), and eggs (64.5–58.3 pg/g). In all of these food groups, apredominance of the tetra- and pentaBDE homologs, followed by hexaBDE, was observed in the sumtotal PBDEs. By contrast, PBDEs were not detected in the groups of fruits, cereals, or tubers. Four typesof commercial fish oils sold in Sweden were found to contain PBDEs (0.2–28.1 ng/g lipid weight)(Haglund et al. 1997). The highest concentration of PBDEs was found in the cod liver oil. These oilswere from products marketed as dietary supplements for humans. The concentrations of PBDEs in seafood from the Inland Sea of Japan were determined for samples collected in 1998 (Hori et al. 2000). BDE 28, BDE 47, BDE 66, BDE 99, BDE 100, BDE 153, and BDE 154 were detected in all analyzed seafood samples. BDE 47 was detected as the predominant congener, with concentrations ranging from58 to 2,100 pg/g wet weight. Harrad et al. (2004) determined the concentrations of several PBDEcongeners in omnivorous and vegan diet samples from the United Kingdom. Median concentrations ofBDE 47, BDE 99, BDE 100, BDE 153, and BDE 154 in omnivorous diet samples were 66.8, 63.8, 10, 20, and 20 pg/g dry weight, respectively. In vegetarian samples, median concentrations of BDE 47, BDE 99, BDE 100, BDE 153, and BDE 154 were 47.2, 56.7, 10.0, 20, and 20 pg/g dry weight, respectively.Concentrations of BDE 47, BDE 99, and total PBDE were found to be statistically higher in omnivorousdiet samples compared to vegetarian diet samples. PBDEs 3956. POTENTIAL FOR HUMAN EXPOSUREBiosolids and Effluents. The concentrations of PBDEs in biosolids (sewage sludge) and effluents are summarized in Table 6-5. PBDEs were detected in 11 biosolids obtained from Virginia, New York, Maryland, and California (Hale et al. 2001c). The total concentrations of pentaBDE in biosolids ranged from 1,100 to 2,290 ng/g dry weight. The concentration of decaBDE (BDE 209) varied widely among biosolids from the four states; the concentration of BDE 209 ranged from 84.8 to 4,890 ng/g dry weight inthe biosolid samples.Levels of PBDEs in biosolids were analyzed using samples from the EPA 2001 National SewageSludge Survey (NSSS) (Venkatesan and Halden 2014). Thirty-two PBDEs were detected in the samples analyzed. The total mean±standard deviation PBDE concentration detected in biosolids composites was9,388±7,778 µg/kg dry weight. Deca, nona, and penta BDE congeners accounted for roughly 57, 18, and 13% of the total, respectively. Using these data and the estimated annual biosolids production and disposal figures in the United States, the annual mean loading rate of PBDEs was estimated to range from47,900 to 60,100 kg/year. Analysis of samples collected between August 2006 and March 2007 (2– 3 years after the voluntary phase-out of penta and octa PBDE formulations) indicated that the levels of8 out of 11 major congeners in biosolids had declined approximately 10–57 % when compared to 2001 levels.Sewage sludge in the vicinity of the Dan River (Virginia) were collected and analyzed for PBDEs (Hale et al. 2002). Congener patterns suggestive of both penta- and decaBDE commercial products were present at concentrations of 1,370 ng/g dry weight (sum of BDE 47 to BDE 154) and from 1,470 ng/g dry weight, respectively. While no known industrial source of pentaBDE discharged to this plant, thedistribution pattern for lower-brominated congeners matched the pentaBDE commercial product. Sewage sludge samples from 13 WWTPs in Germany were sampled (Hagenmaier et al. 1992). The meanconcentration of tri- to heptaBDEs was 8.37 ng/g, with tri-, tetra-, penta-, hexa-, and heptaBDEs atconcentrations of 0.65, 3.06, 3.02, 0.49, and 0.22 ng/g, respectively. Concentrations of penta- and hexaBDEs were highest in these samples. de Boer et al. (2000b) determined the concentration of PBDEsin sewage treatment plant effluents from the Netherlands. The concentration of total PBDEs (the sum ofBDE 47, BDE 99, and BDE 100) ranged from 11 to 35 ng/g dry weight with the overwhelming majority as BDE 47, while the concentrations of BDE 209 ranged from 310 to 920 ng/g dry weight. Kohler et al. (2003) determined the concentrations of decaBDE in sewage sludge from Switzerland between 1993 and 2002. These authors reported that the average concentration of decaBDE increased with time from 220 to 1,100 ng/g dry weight, corresponding to an average increase of 560%. PBDEs 3966. POTENTIAL FOR HUMAN EXPOSURETable 6-5. Concentrations (ng/g Dry Weight) of Several Polybrominated DiphenylEthers (PBDEs) in Biosolids (Sewage Sludge) and EffluentsSample type Location BDE 47 BDE 99 BDE 100 ΣPBDEsa BDE 209 ReferenceSewage sludgeDan River,VirginiaNo data No data No data 2,840* 1,470 Hale et al. 2002Sewage sludge11 biosolidsamples fromVirginia, NewYork,Maryland,and CaliforniaNo data No data No data 1,100– 2,290*No data Hale et al. 2001cSewage sludgeGothenburg,Sweden15 19 3.5 38 No data Nylund et al.1992Sewage sludgeKlippan,Sweden22 18 5.4 45.4 No data Sellström et al.1999; Sellströmand Jansson 1995Sewage SludgeRimbo,Sweden53 53 13 119 No data Sellström et al.1999; Sellströmand Jansson 1995Sewage sludgeThree plants,Stockholm,Sweden39–91 48–120 11–28 98–239 140–350 Sellström et al.1999Sewage sludgeGermany No data No data No data 04–15* No data Hagenmaier etal. 1992Sewage treatmentNetherlands,several sites11–35 <1 No data 11–35 310–920 de Boer et al.2000bplant effluentsaΣPBDEs is the sum of BDE 47, BDE 99, and BDE 100, but if more congeners are included, this is marked with an asterisk (*).BDE = brominated diphenyl ether PBDEs 3976. POTENTIAL FOR HUMAN EXPOSUREWorld Trade Center Site. In 2001, PBDEs were detected in dust and smoke samples taken near theWorld Trade Center disaster site (Lioy 2002). The highest concentration was for decaBDE (i.e., BDE 209), which was present in thermoplastics (e.g., computers). Concentrations of PBDE congeners were 107–174 μg/kg dry weight basis for BDE 47, 51.1–74.1 μg/kg dry weight basis for BDE 100, 155–293 μg/kg dry weight basis for BDE 99, 42.0–53.5 μg/kg dry weight basis for BDE 153, 219–305 μg/kg dry weight basis for BDE 154, and 1,330–2,660 μg/kg dry weight basis for decaBDE (BDE 209). Concentrations of PBDEs were found to be similar to concentrations found in sewage sludge (Lioy 2002). Freshwater Fish. Monitoring data indicated that the concentrations of PBDEs were historicallyincreasing in freshwater organisms, with higher concentrations near point sources. The congener profilesshow the highest concentrations for BDE 47. The presence of PBDEs in freshwater aquatic organisms taken from remote regions suggests that diffuse sources of PBDEs are also important. A sampling ofconcentrations of PBDEs in freshwater fish samples in the United States are summarized in Table 6-6. Fish were sampled from two U.S. lakes, Hadley Lake, Indiana near a possible PBDE point source, and Lake of the Ozarks, Missouri, with no known sources (Dodder et al. 2000). Mean total PBDEconcentrations (sum of BDE 47, BDE 99, BDE 100, BDE 153, and BDE 154) were higher in crappie(Pomoxis annularis) and bluegill (Lepomis macrochirus) from Hadley Lake (1,500 and 1,900 ng/g lipidweight, respectively) than from Lake of the Ozarks (340 and 390 ng/g lipid, respectively). BDE 47, BDE 99, BDE 153, and BDE 154 were the primary congeners. From the Lake of Ozarks, BDE 47 was the dominant congener in fish. The total PBDE concentrations in smelt (Osmerus mordax) from Lakes Superior and Ontario were 150±9 and 240±30 ng/g lipid, respectively (Dodder et al. 2002). Thedominant congeners in these fish were BDE 47 and BDE 99. An analysis of fish tissue samples from selected locations in Washington State showed that total PBDEconcentrations ranged from 29 ng/g lipid in rainbow trout from a remote spring-fed stream (Douglas Creek, Washington) to 19,000 ng/g lipid in rainbow trout from the urbanized Spokane River, Washington (Johnson and Olson 2001). The tetra- and pentaBDE isomers were the major compounds present. TetraBDE to hexaBDE were found in carp (Cyprinus carpio) from the Buffalo River (New York), apolluted area around the Great Lakes (Loganathan et al. 1995). TetraBDEs dominated the congenerpattern with 94–96% of total PBDEs. TetraBDE and pentaBDE concentrations ranged from 13 to 22 ng/g fresh weight. Asplund et al. (1999a) found tri- to hexaBDEs in steelhead trout (Oncorhynchus mykiss) sampled in 1995 from Lake Michigan. The combined concentration of BDE 47, BDE 99, BDE 100, PBDEs 3986. POTENTIAL FOR HUMAN EXPOSURETable 6-6. Concentrations (ng/g Lipid Weight, Except as Noted) of SeveralPolybrominated Diphenyl Ethers (PBDEs) in Freshwater Fish Samplesfrom the United StatesSample type Location BDE 47 BDE 99 BDE 100 ΣPBDEsa BDE 209 ReferenceLake trout Lake Michigan 93–230 9.0–48 12–45 120–350* No data Streets et al. 2006Lake trout Lake Superior 22–79 8.5–53 5.2–19 39–180* No data Streets et al. 2006Lake trout Lake Huron 32–59 7.8–13 6.5–12 50–94* No data Streets et al. 2006Lake trout Lake Ontario 45–140 6.5–34 7.7–24 64–230* No data Streets et al. 2006Lake Trout Lake Ontario 4.3–114 59–680 7.4–1,285 269–3,339* 2.3–12 Ismail et al.2009Alewife Grand Traverse 16 No data No data 36 No data Stapleton andBay, Lake Baker 2003MichiganBloater Grand Traverse 11 (fw) No data No data 23 (fw) No data Stapleton andchub Bay, Lake Baker 2003MichiganBluegill Hadley Lake, 420 320 240 1,900 No data Dodder et al.Indiana 2000Bluegill Lake of the 200 91 59 390 No data Dodder et al.Ozarks, Missouri 2000Burbot Grand Traverse 43 (fw) No data No data 86 (fw) No data Stapleton andBay, Lake Baker 2003MichiganCarp United States No data No data No data 13–22* No data Loganathan et(fw) al. 1995Carp Detroit River, 3.0 (fw) 0.50 (fw) 0.48 (fw) 40.7* No data Rice et al. 2002Grosse Isle,MichiganCarp Des Plaines 2.54 (fw) 0.5 (fw) 0.44 (fw) 281* No data Rice et al. 2002River, Joliet,IllinoisCarp Des Plaines 1.34 (fw) 0.50 (fw) 0.49 (fw) 78.3* No data Rice et al. 2002River, Joliet,IllinoisCarp (fillet) Yakima River, No data No data No data 22 (fw) No data Johnson and Washington Olson 2001Crappie Hadley Lake, 250 430 150 1,500* No data Dodder et al.Indiana 2000Crappie Lake of the 190 78 59 340* No data Dodder et al.Ozarks, Missouri 2000Deepwater Grand Traverse 2.8 (fw) No data No data 3 (fw) No data Stapleton andsculpin Bay, Lake Baker 2003MichiganLake trout Grand Traverse 75 (fw) No data No data 126 (fw) No data Stapleton andBay, Lake Baker 2003Michigan PBDEs 3996. POTENTIAL FOR HUMAN EXPOSURETable 6-6. Concentrations (ng/g Lipid Weight, Except as Noted) of SeveralPolybrominated Diphenyl Ethers (PBDEs) in Freshwater Fish Samplesfrom the United StatesSample type Location BDE 47 BDE 99 BDE 100 ΣPBDEsa BDE 209 ReferenceLake trout Lake Ontario, No data No data No data 540* No data Alaee et al. 1999United StatesLake trout Lake Ontario, 58 (fw) 14 (fw) 5.7 (fw) No data No data Luross et al.United States 2002Lake trout Lake Huron, No data No data No data 240* No data Alaee et al. 1999United StatesLake trout Lake Huron, 27 (fw) 7.7 (fw) 3.8 (fw) No data No data Luross et al.United States 2002Lake trout Lake Superior, No data No data No data 140* No data Alaee et al. 1999United StatesLake trout Lake Superior, 29 (fw) 12 (fw) 4.1 (fw) No data No data Luross et al.United States 2002Lake trout Lake Erie, No data No data No data 117* No data Alaee et al. 1999United StatesLake trout Lake Erie, 16 (fw) 2.0 (fw) 2.5 (fw) No data No data Luross et al.United States 2002Largescale Yakima River, No data No data No data 64 (fw) No data Johnson and sucker Washington Olson 2001(whole)Largescale Spokane River, No data No data No data 105 (fw) No data Johnson and sucker Washington Olson 2001(whole)Mountain Spokane River, No data No data No data 1,250 (fw) No data Johnson and whitefish Washington Olson 2001(whole)Rainbow Douglas Creek, No data No data No data 1.5 (fw) No data Johnson and trout Washington Olson 2001(whole)Rainbow Spokane River, No data No data No data 20–174 No data Johnson and trout Washington (fw) (fillet) Olson 2001297 (fw)(whole)Northern Spokane River, 59–160 0.3–<0.4 17–47 No data Not Furl and pike Washington detected Meredith 2010minnowMountain Spokane River, 127–942.6 81–942.6 26.3– No data Not Furl and whitefish Washington 368.1 detected Meredith 2010Largescale Spokane River, 87–270 0.2–<4.4 13–45 No data Not Furl and sucker Washington detected Meredith 2010Salmon Grand Traverse 34 (fw) No data No data 95 (fw) No data Stapleton andBay, Lake Baker 2003MichiganSalmon Lake Michigan, 52.1 (fw) 9.3 (fw) 9.7 (fw) 2,440 No data Manchester-United States Neesvig et al.2001Smelt Lake Superior, 5.7 (fw) 1.8 (fw) 0.98 (fw) 150 <1.5 (fw) Dodder et al.United States 2002 PBDEs 4006. POTENTIAL FOR HUMAN EXPOSURETable 6-6. Concentrations (ng/g Lipid Weight, Except as Noted) of SeveralPolybrominated Diphenyl Ethers (PBDEs) in Freshwater Fish Samplesfrom the United StatesSample type Location BDE 47 BDE 99 BDE 100 ΣPBDEsa BDE 209 ReferenceSmelt Lake Ontario, 10 (fw) 5.3 (fw) 1.6 (fw) 240 <1.6 (fw) Dodder et al.United States 2002Starry Columbia River, No data No data No data 30 (fw) No data Johnson and flounder Washington Olson 2001(whole)Steelhead Lake Michigan, 1,700 600 360 3,000* No data Asplund et al.trout United States 1999bWhitefish Columbia River, No data No data No data 72 (fw) No data Rayne et al.United States 2003aWhitefish Grand Traverse 9.8 (fw) No data No data 18 (fw) No data Stapleton andBay, Lake Baker 2003MichiganaΣPBDEs is the sum of BDE 47, BDE 99, and BDE 100, but if more congeners are included, this is marked with an asterisk (*).BDE = brominated diphenyl ether; dw = dry weight; fw = fresh weight PBDEs 4016. POTENTIAL FOR HUMAN EXPOSUREBDE 153, and BDE 154 was 3,000 ng/g lipid weight (Asplund et al. 1999b). Lake trout (Salvelinusnamaycush) from Lakes Ontario, Huron, and Superior were also found to have di- to heptaBDEs withcombined concentrations of 545, 237, and 135 ng/g lipid weight, respectively (Alaee et al. 1999). Lake trout from Lake Erie had 117 ng/g lipid weight (Luross et al. 2000). Variations in local sources, combined with atmospheric transport, may explain differences that were seen in congener profiles for thedifferent lakes. A retrospective temporal study for the years 1978, 1983, 1988, 1993, and 1998 usingarchived trout samples from Lake Ontario show a dramatic increase in total PBDE concentrations overtime (Luross et al. 2000). At 50 freshwater sites in Virginia, muscle samples from 253 fish samples were collected and analyzed for PBDEs (Hale et al. 2000, 2001b). Approximately 85% of the samplescontained BDE 47, the predominant congener, at measurable concentrations. Concentrations were>1,000 ng/g lipid weight at 9 of 50 sites. The highest combined PBDE concentrations (up to 57,000 ng/g lipid weight) were observed in carp downstream of textile and furniture facilities. BDE 47 concentrationswere greater than 2,2’,4,4’,5,5’-hexachlorobiphenyl (PCB 153) concentrations in 58% of the samplesanalyzed. PBDEs were identified in fish collected from the Detroit River (Michigan) and Des Plaines Rivers (Illinois). In Detroit River fish (carp and largemouth bass), the congener patterns were dominatedby BDE 47; however, in the Des Plaines River carp, the dominant congeners were heptaBDE congeners (BDE 181 and BDE 183), lesser amounts of BDE 190, and two hexaBDEs (BDE 154 and BDE 153). Possible sources for the heptaBDE congeners were not obvious Ozarks (340 and 390 ng/g lipid,respectively). BDE 47, BDE 99, BDE 153, and BDE 154 were the primary congeners. From the Lake ofOzarks, BDE 47 was the dominant congener in fish. The total PBDE concentrations in smelt (O. mordax) from Lakes Superior and Ontario were 150±9 and 240±30 ng/g lipid, respectively (Dodder et al. 2002). The dominant congeners in these fish were BDE 47 and BDE 99. An analysis of fish tissue samples fromselected locations in Washington State showed that total PBDE concentrations ranged from 29 ng/g lipid in rainbow trout from a remote spring-fed stream (Douglas Creek, Washington) to 19,000 ng/g lipid inrainbow trout from the urbanized Spokane River, Washington (Johnson and Olson 2001). The tetra- and pentaBDE isomers were the major compounds present. TetraBDE to hexaBDE were found in carp (C. carpio) from the Buffalo River (New York), a polluted area around the Great Lakes (Loganathan et al. 1995). TetraBDEs dominated the congener pattern with 94–96% of total PBDEs. TetraBDE and pentaBDE concentrations ranged from 13 to 22 ng/g fresh weight. Asplund et al. (1999a) found tri- to hexaBDEs in steelhead trout (O. mykiss) sampled in 1995 from Lake Michigan. The combined concentration of BDE 47, BDE 99, BDE 100, BDE 153, and BDE 154 was 3,000 ng/g lipid weight(Asplund et al. 1999b). Lake trout (S. namaycush) from Lakes Ontario, Huron, and Superior were also PBDEs 4026. POTENTIAL FOR HUMAN EXPOSUREfound to have di- to heptaBDEs with combined concentrations of 545, 237, and 135 ng/g lipid weight,respectively (Alaee et al. 1999).The National Study of Chemical Residues in Lake Fish Tissue (1998–2009) is one of the statistically-based surveys conducted by EPA that analyzed the concentrations of PBDEs and other contaminants infish from 500 lakes in the continental United States (EPA 2009c, 2013k). The most prevalent PBDEcongeners detected in both predator and bottom dweller fish were reported as BDE 47, BDE 99, andBDE 100 (EPA 2013k). Ismail et al. (2009) studied the temporal trends of PBDE congeners in trout obtained from Lake Ontario from 1979 to 2004. Concentrations of most PBDE congeners (BDE 28, BDE 47, BDE 99, BDE 100, BDE 153, and BDE 154) increased dramatically from 1979 to the mid-1990s and then either leveled offor decreased from 1998 to 2004; average concentration and standard error values are presented inFigure 6-2. The temporal trend of BDE 209 was different than for the other congeners, however. Concentrations of BDE 209 in lake trout increased more slowly from 1979 until the mid-1990s, but itsconcentrations increased dramatically from the late 1990s (3.3±0.8 ng/g in 1998) until 2004 (12±5.3 ng/gin 2004), corresponding to its increased use in consumer products.The concentrations of PBDEs in freshwater fish samples from Europe are summarized in Table 6-7. Between 1986 and 1988, concentrations of BDE 47, BDE 99, and BDE 100 were measured in whitefish(Coregonus spp.) from a remote mountain lake in Northern Sweden (Lake Storvindeln), in Arctic char(Salvelinus alpinus) from a heavily populated lake (Lake Vättern) in south-central Sweden with numerousmunicipal and industrial point sources, and in trout (Salmo trutta) and pike (Esox lucius) from severalsites along Dalslands Canal in west central Sweden (Jansson et al. 1993). No point sources of PBDEswere identified from these sites. Whitefish from the remote lake contained the lowest concentrations(26 ng/g lipid weight) of PBDEs, whereas the Arctic char, from a heavily populated lake, contained 520 ng/g lipid weight PBDEs. In both samples, BDE 47 was the predominant congener. PBDEconcentration in pike and trout from the Dalslands Canal ranged from 180 to 210 ng/g lipid weight andfrom 280 to 1,200 ng/g lipid weight, respectively. The congener pattern in these samples was similar tothe technical mixture, Bromkal 70-5DE, with equal quantities of both BDE 47 and BDE 99. Theconcentrations in pike and trout are of the same order of magnitude as in the Arctic char, indicating thespread of PBDEs from diffuse sources (de Wit 2002). In 1979 and 1980, high concentrations of tri- to hexaBDEs (range, 950–27,000 ng/g lipid weight in muscle tissues) were measured in fish sampled alonga river in Sweden (Viskan) where numerous textile industries are located (Andersson and Blomkvist PBDEs 4036. POTENTIAL FOR HUMAN EXPOSUREFigure 6-2. Temporal Trends in the Concentrations (ng/g Lipid Weight) of PBDECongeners in Lake Trout from Lake Ontario, CanadaSource: Ismail et al. 2009 PBDEs 4046. POTENTIAL FOR HUMAN EXPOSURETable 6-7. Concentrations (ng/g Lipid Weight Except as Noted) of SeveralPolybrominated Diphenyl Ethers (PBDEs) in Freshwater Fish Samples from EuropeSample type Location BDE 47 BDE 99 BDE 100 ΣPBDEsa BDE 209 ReferenceFreshwater Pyrenees 0.40–0.51 0.24 0.16–0.18 No data No data Gallego et al.fish Mountains 2007Freshwater Tatras 0.20–0.26 0.17–0.21 0.043 No data No data Gallego et al.fish Mountains 2007Arctic char Lake Vättern, 400 64 51 520 No data Sellström et al.Sweden 1993Bream Netherlands 0.2–130 Not No data No data No data de Boer et al.(several sites) (dw) detected 2000bEels Netherlands <20–1,400 No data No data <50–1,700 No data de Boer 1990Osprey Sweden 1,800 140 200 2,140 No data Sellström et al.1993Pike Dalslands canal, 94–98 60–79 25–36 180–210 No data Sellström et al.Sweden 1993Pike River Viskan, <46–2,000 <37– <14– <130– Trace Sellström et al.Sweden, 1,600 1,000 4,600 1998aupstream and downstreamSeveral Germany No data No data No data 19–983* No data Krüger 1988fishspeciesTrout Dalslands canal, 120–460 130–590 33–150 280–1,200 No data Sellström et al.Sweden 1993Whitefish Lake 15 7.2 3.9 26 No data Sellström et al.Storvindeln, 1993SwedenWhitefish Lake Geneva, 26 13 2.5 44* No data Zennegg et al.Switzerland 2003Whitefish Lake Greifen, 96 52 9.1 165* No data Zennegg et al.Switzerland 2003Whitefish Lake Biel, 75.9 39 7.1 128* No data Zennegg et al.Switzerland 2003Whitefish Lake Lucerne, 56 46 10 121* No data Zennegg et al.Switzerland 2003Whitefish Lake Zürich, 56 25 4.5 89* No data Zennegg et al.Switzerland 2003Whitefish Lake Nauchatel 41 20 4.0 68* No data Zennegg et al.Switzerland 2003Whitefish Lake Constance, 32 15 2.9 52* No data Zennegg et al.Switzerland 2003 PBDEs 4056. POTENTIAL FOR HUMAN EXPOSURETable 6-7. Concentrations (ng/g Lipid Weight Except as Noted) of SeveralPolybrominated Diphenyl Ethers (PBDEs) in Freshwater Fish Samples from EuropeSample type Location BDE 47 BDE 99 BDE 100 ΣPBDEsa BDE 209 ReferenceWhitefish Lake Thun,Switzerland19 12 2.5 36* No data Zennegg et al.2003aΣPBDEs is the sum of BDE 47, BDE 99, and BDE 100, but if more congeners are included, this is marked with an asterisk (*).BDE = brominated diphenyl ether PBDEs 4066. POTENTIAL FOR HUMAN EXPOSURE1981). These textile industries have used PBDEs in the production of textiles. BDE 47 was the predominant congener at 70–80% of the total PBDEs. In 1977, the PBDEs were not detected in fish sampled at the same sites. The elevated concentrations of BDE 47, BDE 99, and BDE 100 were laterconfirmed in a follow-up study in which fish were caught from approximately the same locations (Sellström et al. 1993). In the current study, BDE 47 was the predominant congener at 65–96% of totalPBDEs. Several fish species were sampled (pike, perch, bream, eel, tench, and sea trout) in these studies.In 1995, fresh samples of pike and sediments were collected at four of eight sites along River Viskan inorder to search for point sources of contaminants. The combined concentrations of BDE 47, BDE 99, and BDE 100 ranged from not detected to 4,600 ng/g lipid weight; with BDE 47 being the predominantcongener (50–90% of total). DecaBDE (BDE 209) was found in a few fish at trace amounts. The lowestconcentrations of the PBDEs were found upstream of the industries. The concentrations of PBDEsincreased further downstream as more industries were passed (Sellström et al. 1998a). Concentrations ofBDE 47 ranged from <20 to 1,700 ng/g lipid in eels (Anguilla anguilla) from Dutch rivers and lakes (at10 locations); BDE 47 comprised 70% of the total PBDEs (de Boer 1990). Bream (Abramis brama) sampled from several sites in the Netherlands had concentrations of BDE 47 ranging from 0.2 to 130 ng/g dry weight (de Boer et al. 2000b). BDE 99 was below the detection limits. BDE 153 ranged from<0.04 to 4.1 ng/g dry weight. Allchin et al. (1999) conducted a study of PBDEs in plaice (Pleuronectes platessa), flounder (Platichthys flesus), and dab (Limanda limanda) collected in the estuaries of rivers inthe United Kingdom. Suspected sources of PBDEs in the estuaries include a manufacturer of pentaBDEand octaBDE, several industries using pentaBDE, and several landfills receiving wastes suspected tocontain PBDEs. Concentrations of BDE 47, BDE 99, pentaBDE (as technical mixture DE-71), and octaBDE (as technical mixture DE-79) in fish ranged from not detected to 9,500 ng/g lipid weight, notdetected to 370 ng/g lipid weight, 47–1,200 ng/g lipid weight, and not detected to 1,200 ng/g lipid weight.The highest concentrations were at Tees Bay downstream from a manufacturing plant on the River Tees.These results are similar to the situation found in Sweden along the River Viskan (Andersson and Blomkvist 1981; Sellström et al. 1993). Freshwater mussels (Dreissena polymorpha) were collected atseveral locations in the Netherlands and analyzed for BDE 47, BDE 99, BDE 153, and BDE 209 (de Boeret al. 2000b). Concentration ranges for the congeners were 0.7–17, 0.4–11, and <0.1–1.5 ng/g dry weightfor BDE 47, BDE 99, and BDE 153, respectively; BDE 209 was below the detection limit. Poma et al.(2014) analyzed freshwater zebra mussels in Lake Maggiore, Northern Italy for the presence of PBDEs and other brominated flame retardants. Total tri- to heptaBDE concentrations ranged from 1.0 ng/g forsamples obtained in in May 2011 to 144.6 ng/g for samples collected in September 2012. The authorsnoted that even though penta- BDE was banned in Europe in 2004, increasing concentrations of tri- to heptaBDE congeners in mussels from 2011 to 2012 were observed in the samples obtained. Average PBDEs 4076. POTENTIAL FOR HUMAN EXPOSUREvalues of hepta to decaBDE congeners in zebra mussels ranged from 88.2 to 182.8 ng/g and tended to bedominated by BDE 209, with average concentrations of 71.2–144.7 ng/g (Poma et al. 2014). Thepresence of the lower brominated octa and hepta congeners was likely due to metabolism of BDE 209 within the organism or the environmental debromination of BDE 209 followed by uptake by the zebramussels.Saltwater Fish. Spatial trends show higher concentrations of lower-brominated BDE congeners foundnear human populated areas. The congener profiles show the highest concentrations for BDE 47. Representative concentrations of several PBDEs in marine aquatic species are summarized in Table 6-8. BDEs were detected in 10/10 white croakers and 8/8 shiner surfperch obtained from the San Francisco Bay at concentrations of (total PBDEs) 470–2, 260 and 730–3,930 ng/g lipid, respectively (Klosterhaus etal. 2012). Mean levels of total PBDEs in halibut, jack smelt, leopard shark, northern anchovy, shinersurfperch, striped bass white croaker, and white sturgeon collected from the San Francisco Bay in 2009 ranged from 1.5 to 8.3 ng/g wet weight (Sutton et al. 2015). In the year 2000, sole liver collected fromfive sites along the Canadian west coast (Crofton, Bamfield, Kitimat, Trincomali, and Vancouver) were analyzed for 14 BDE congeners (Ikonomou et al. 2002); the total PBDE concentrations were 64–340 ng/g lipid while the three highest congener concentrations were 27–160 ng/g lipid (BDE 47), 8.5–54 ng/g lipid(BDE 100), and 9.5–46 ng/g lipid (BDE 99). The highest concentrations were found in sole samplescollected near Vancouver, Canada. DecaBDE was not detected in these samples at the level of proceduralblank. Farmed salmon collected at two locations in Canada were analyzed for PBDE congeners (Easton et al. 2002). Forty-one congeners were detected with BDE 47 at the highest concentration (690 and 2,600 ng/g wet weight) followed by BDE 99 and BDE 100; total BDE congener concentrations were1,188 and 4,147 ng/g wet weight for the two samples. Likewise, wild salmon from four locations in Canada were analyzed for BDE congeners. Concentrations were a factor of 10 lower for these samples compared to farmed salmon samples. The total PBDE concentration for the 41 detected congeners ranged from 38.7 to 485.2 ng/g wet weight. The concentration of the highest congener, BDE 47, ranged from29 to 280 ng/g wet weight (Easton et al. 2002). PBDE concentrations in skipjack tuna from Asian offshore waters, off-Seychelles, off-Brazil, and open seas were determined for samples collected during1996–2001 (Ueno et al. 2003). The concentration of total BDEs in muscles tissues ranged from notdetected (<0.05 ng/g lipid) to 53 ng/g lipid. The concentration of the highest congener in muscle tissues, BDE 47, ranged from <0.1 to 15 ng/g lipid. BDE 99, BDE 100, BDE 153, and BDE 154 also weredetected; BDE 209 was below the detection limit (<5.0 ng/g lipid) for these samples. Samples collectedoff the coast of the Seychelles (relatively pristine area) did not have detectable concentrations of anyPBDEs, while samples collected in industrial areas of southeast Asia had the highest. Fall-caught PBDEs 4086. POTENTIAL FOR HUMAN EXPOSURETable 6-8. Concentrations (ng/g Lipid Weight) of Several Polybrominated Diphenyl Ethers (PBDEs) in Marine Aquatic SpeciesSample type Location BDE 47 BDE 99 BDE 100 ΣPBDEsa BDE 209 ReferenceCalifornia San No data No data No data 1.8±0.5* ND Sutton et al.2015halibut Francisco Bay(ww)Jack smelt San No data No data No data 1.4±0.4* ND Sutton et al.2015Francisco Bay(ww)Leopard sharkSan Francisco BayNo data No data No data 5.0±1.2* (ww)ND Sutton et al.2015Northern San No data No data No data 7.9±2.9* ND Sutton et al.2015anchovy Francisco Bay(ww)Shiner San No data No data No data 8.3±2.9* ND Sutton et al.2015surfperch Francisco Bay(ww)Striped bass San Francisco BayNo data No data No data 5.0±2.6* (ww)ND Sutton et al.2015White San No data No data No data 4.3±2.5* ND Sutton et al.2015croacker Francisco Bay(ww)White San No data No data No data 2.8±1.3* ND Sutton et al.2015sturgeon Francisco Bay(ww)Winter Northwest 35 2.5 6.4 52* ND Shaw et al. 2009Flounder AtlanticAtlantic Northwest 40 6.9 6.8 82* ND Shaw et al. 2009Herring AtlanticAmerican Northwest 42 4.0 7.0 69* 1.9 Shaw et al. 2009Plaice AtlanticWhite Hake Northwest 25 0.63 7.2 42* 0.91 Shaw et al. 2009AtlanticAlewife Northwest 8.3 3.6 1.7 18* ND Shaw et al. 2009AtlanticAtlantic Northwest 20 7.5 4.1 69* 1.6 Shaw et al. 2009Mackerel AtlanticSilver Hake Northwest 19 6.3 4.0 38* ND Shaw et al. 2009AtlanticFarmed SalmonCanada 690;2,600 (ww)140; 390(ww)130; 470(ww)1,187; No data4,147 (ww)Easton et al. 2002Salmon Canada 29–280 ND–97 4.2–43 38.7–485.2 No data Easton et al. 2002(wild) (ww) (ww) (ww) (ww) PBDEs 4096. POTENTIAL FOR HUMAN EXPOSURETable 6-8. Concentrations (ng/g Lipid Weight) of Several Polybrominated Diphenyl Ethers (PBDEs) in Marine Aquatic SpeciesSample type Location BDE 47 BDE 99 BDE 100 ΣPBDEsa BDE 209 ReferenceSole liver West coast, 27–160 9.5–46 8.5–54 64–340* ND Ikonomou et al.Canada 2002Skipjack Seychelles, <0.1 <0.05 <0.05 ND <5.0 Ueno et al. 2003tuna Indian OceanSkipjack East China 9.0–15 2.4–4.7 3.4–4.4 23–34 <5.0 Ueno et al. 2003tuna SeaSkipjack Pacific 2.9–7.9 0.18–3.0 0.56–2.1 5.8–21 <5.0 Ueno et al. 2003tuna OceanHerring Baltic Sea 19–38 7.8–17 3.4–6 30–61 No data de Wit 2002 Sellström et al.1993Herring Baltic Sea 3.2–27 ND–2.9 1.3–1.9 3.2–32 No data Haglund et al. 1997Herring Baltic Sea 7.6–24 4.3–3.9 No data 12.9–28.3* No data Strandman et al.1999Herring Baltic Sea 6.3 0.6 0.8 12* No data Burreau et al. 1999Herring Kattegatt, 12 3.4 1.6 17 No data de Wit 2002;Sweden Sellström et al.1993Herring North Sea 8.4–100 No data No data No data No data de Boer 1990Sprat Baltic Sea 17.5– 1.9–9.5 No data 21–149* No data Strandman et al.(different 140.8 1999age groups)Sprat Baltic Sea 4.3 0.7 0.8 8.4* No data Burreau et al. 1999Cod liver North Sea 170 No data No data 1.9–360 No data de Boer 1989Salmon Baltic Sea 167 52 44 220 No data Haglund et al. 1997Salmon Baltic Sea 190 52 46 290 No data Asplund et al. 1999bSalmon Baltic Sea 46 7.3 6.4 86* No data Burreau et al. 1999Several fish Japan No data No data No data 0.1–17* No data Watanabe et al.species 1987Yellowfin Japan 0.5 0.4 0.25 1.9* No data Ohta et al. 2000tunaYellowtail Japan 17 4.5 4.0 30.5* No data Ohta et al. 2000Yellowtail Japan 29 3.3 5.3 44* No data Ohta et al. 2000(cultured)Salmon Japan 22 8.1 5.3 46* No data Ohta et al. 2000 PBDEs 4106. POTENTIAL FOR HUMAN EXPOSURETable 6-8. Concentrations (ng/g Lipid Weight) of Several Polybrominated Diphenyl Ethers (PBDEs) in Marine Aquatic SpeciesSample type Location BDE 47 BDE 99 BDE 100 ΣPBDEsa BDE 209 ReferenceSeveralflatfishSeven riverestuaries,Great Britain73–9,500 16–790 No data No data ND Allchin et al. 1999Flounder Netherlands,several sites0.6–20 (dw)<0.01–4.6 No data No data No data de Boer et al.2000baΣPBDEs is the sum of BDE 47, BDE 99, and BDE 100, but if more congeners are included, this is marked with an asterisk (*).BDE = brominated diphenyl ether; dw = dry weight; ND = not detected; ww = wet weight PBDEs 4116. POTENTIAL FOR HUMAN EXPOSUREherring (Clupea harengus) muscle from five sites along the Swedish coast was analyzed for BDE 47, BDE 99, and BDE 100; the combined concentration of these three congeners ranged from 17 to 61 ng/g lipid, with BDE 47 being the dominant congener (Sellström et al. 1993). Likewise, the concentration ofBDE 47 in Baltic herring ranged from 3.2 to 27 ng/g lipid in different age groups; the combined concentration of BDE 47, BDE 99, and BDE 100 ranged from 3.2 to 32 ng/g lipid (Haglund et al. 1997);2-year-old herring had the lowest concentrations and 5-year-old herring had the highest concentrations. Similarly, Strandman et al. (1999) observed increasing concentrations with age of BDE 47, BDE 99, and BDE 153 in Baltic sprat (Sprattus sprattus, age 3–13 years). However, this trend was not evident forherring. BDE 47 was the primary congener with concentrations ranging from 7.6 to 24 ng/g lipid weight for 1–3-year-old sprat, 17–140 ng/g lipid weight for 3–13-year-old sprat, and 7.6–24 ng/g lipid weight inherring. The concentrations of BDE 47, BDE 99, and BDE 100 in whole-body composites of herringwere 6.21, 0.62, and 0.81 ng/g lipid, respectively; in sprat, the concentrations were 4.32, 0.71, and 0.80 ng/g lipid, respectively (Burreau et al. 1999). Baltic sea herring had similar concentrations ofBDE 47 (46.3 ng/g lipid) compared to 8.4–100 ng/g lipid of BDE 47 found by de Boer (1990) for herringcollected from three regions of the North Sea. BDE 47, BDE 99, and BDE 153 concentrations in Balticsalmon (Salmo salar) muscle were 167, 52, and 4.2 ng/g lipid, respectively (Haglund et al. 1997). BDE 47, BDE 99, and BDE 100 concentrations were 47, 7.2, and 6.3 ng/g lipid, respectively, in whole-body composites (Burreau et al. 1999). In another study, the concentrations of BDE 47, BDE 99, and BDE 100 were determined in muscle, ripe eggs, and blood plasma from Baltic salmon (Asplund et al. 1999a). The mean concentrations of PBDEs in tissues from Baltic salmon (ng/g lipid weight) were asfollows: BDE 47 (muscle, 190; ripe eggs, 64; blood, 190), BDE 99 (muscle, 52; ripe eggs, 16; blood, 55), and BDE 100 (muscle, 46; ripe eggs, 18; blood, 59). Cod (Gadus morhua) liver samples at threelocations of the North Sea had combined concentrations of BDE 47 and BDE 99 of 1.9–360 ng/g lipid (deBoer 1989). BDE concentrations in flounder were 0.6–20 ng/g dry weight for BDE 47 and <0.01– 4.6 ng/g dry weight for BDE 99 from several sites in the Netherlands (de Boer et al. 2000b). Concentrations of BDE 153 and BDE 209 were below the detection limit. In 1996, de Boer et al. (2001)measured the concentrations of two BDE congeners in flounder liver samples from the Amsterdam andRotterdam harbors, and off the Dutch coast; BDE 47 and BDE 99 ranged from 15 to 280 and from <2 to 24 ng/g lipid weight, respectively. Olsson et al. (1999) detected BDE 47 in perch (Perca fluviatilis) fromLatvia in a study examining environmental contamination in coastal areas of the former Soviet Union; theconcentration of BDE 47 ranged from 6.4 to 10 ng/g lipid weight in the perch. Watanabe et al. (1987) detected PBDEs in numerous marine fish and shell fish in Japan. TetraBDE andpentaBDE concentrations ranged from 0.1 and 17 ng/g fresh weight, with tetraBDE being the major PBDEs 4126. POTENTIAL FOR HUMAN EXPOSUREcongener. DecaBDE was also detected in a mussel sample from Osaka Bay (at 1.4 μg/kg wet weight). Japanese market fish were analyzed for PBDEs. The highest combined PBDE concentrations (BDE 28, BDE 47, BDE 66, BDE 99, BDE 100, BDE 153, and BDE 154) were in salmon, cultured yellowtail, and wild yellowtail muscle (46, 44, and 30.5 ng/g lipid weight, respectively) and lowest concentrations inyellowfin tuna (1.9 ng/g lipid weight) (Ohta et al. 2000). BDE 47 was major congener in all samples. Inanother study, several fish species from Japan were analyzed for 15 BDE congeners (Hori et al. 2000). The PBDE concentrations ranged from 0.00136 to 2.1 ng/g fresh weight, with BDE 47 as the predominantcongener. Seven species of marine fish (conger eel, flounder, gray mullet, horse mackerel, red sea bream,sea bass, and yellowtail) were collected from the Inland Seas near Seto, Japan (Akutsu et al. 2001). Seven PBDEs (BDE 28, BDE 47, 2,3’,4,4’-tetraBDE [BDE 66], BDE 99, BDE 100, BDE 153, and BDE 154) were detected in all samples, with BDE 47 being the most abundant congener. Concentrationsof total PBDEs in gray mullets and yellowtails were 63 and 15 ng/g lipid weight, respectively.Marine Aquatic Organisms. Marine mussels (Mytilus edulis) collected at several locations in the Netherlands and analyzed for BDE 47, BDE 99, BDE 153, and decaBDE (BDE 209) (de Boer et al.2000b). Concentrations of BDE 47 and BDE 99 were 0.9–4.3 and 0.3–1.6 ng/g dry weight, respectively. BDE 153 and BDE 209 were not detected. Di- to heptaBDE were analyzed for in hepatopancreas samples from Dungenes crab from several sites on the Strait of Georgia, British Columbia, Canada (Ikonomou et al. 1999). The primary congener detected was BDE 47. The combined concentration ofBDE 47 and BDE 99 was approximately 100–350 ng/g lipid weight.Marine Animals. In marine animals, temporal trends show increasing concentrations of lowerbrominated BDE congeners with higher concentrations found near human-populated areas. In all marineanimal studies, the congener profiles show the highest concentrations for BDE 47. The concentrations ofseveral PBDEs in marine animals are summarized in Table 6-9. Frouin et al. (2011) reported PBDEconcentrations measured from serum and blubber samples obtained from six harbor seal pups (Phocavitulina) live captured in May 2007, six harbor seal pups (Phoca vitulina) live captured in May 2008, six grey seal pups (Halichoerus grypus) live captured early January 2008, and six harp seal pups (Phoca groenlandica) live captured in March 2008 from the Gulf of St. Lawrence or the St. Lawrence Estuary.The ƩPBDEs in serum (BDE 47, BDE 99, BDE 100, BDE 153, BDE 154, and BDE 155) and blubber(BDE 28, BDE 47, BDE 49, BDE 66, BDE 99, BDE 100, BDE 153, BDE 154, BDE 155, and BDE 183)were strongly correlated. BDE 47 was detected in all serum samples and accounted for 66–73% ofƩPBDEs. ƩPBDE concentrations in lipid ranged from 21 to 530 ng/g lipid weight and from 34 to 600 ng/g lipid weight in serum. PBDEs have been detected in several species of seal from several PBDEs 4136. POTENTIAL FOR HUMAN EXPOSURETable 6-9. Concentrations (ng/g Lipid Weight) of Several Polybrominated Diphenyl Ethers (PBDEs) in Marine AnimalsSample type Location BDE 47 BDE 99 BDE 100 ΣPBDEsa BDE 209 ReferenceBottlenose Gulf of Mexico No data No data No data 8,000 No data Kuehl anddolphin Haebler 1995Harbor seal Northwest 904 134 49 1,385* 1.2 Shaw et al. 2009AtlanticHarbor seal San Francisco No data No data No data 530–5,075* No data Klosterhaus et al.Bay, California 2012Cormorant San Francisco No data No data No data 3,425–5,550* No data Klosterhaus et al.eggs Bay, California 2012Harbor seal San Francisco 46– 17–303 No data No data No data She et al. 2000Bay, California 6,682Cormorant Suisan Bay, No data No data No data 19,000±4,000* No data Sutton et al. 2015eggs CA (2002)Cormorant Suisan Bay, No data No data No data 6,100±5,200* No data Sutton et al. 2015eggs California (2006)Cormorant Suisan Bay, No data No data No data 440±1708 No data Sutton et al. 2015eggs California (2009)Cormorant Suisan Bay No data No data No data 1,300±900* No data Sutton et al. 2015eggs California (2012)Cormorant Central Bay, No data No data No data 9,100±2200 No data Sutton et al. 2015eggs California (2002)Cormorant Central Bay, No data No data No data 3,700±500 No data Sutton et al. 2015eggs California (2004)Cormorant Central Bay, No data No data No data 1,800±400 No data Sutton et al. 2015eggs California (2006)Cormorant Central Bay, No data No data No data 1,800±300 No data Sutton et al. 2015eggs California (2009)Cormorant Central Bay, No data No data No data 1,100±200 No data Sutton et al. 2015eggs California (2012)Cormorant South Bay No data No data No data 4,200±400 No data Sutton et al. 2015eggs California (2002)Cormorant South Bay No data No data No data 3,300±300 No data Sutton et al. 2015eggs California (2004) PBDEs 4146. POTENTIAL FOR HUMAN EXPOSURETable 6-9. Concentrations (ng/g Lipid Weight) of Several Polybrominated Diphenyl Ethers (PBDEs) in Marine AnimalsSample type Location BDE 47 BDE 99 BDE 100 ΣPBDEsa BDE 209 ReferenceCormorant South Bay No data No data No data 4,400±2,100 No data Sutton et al. 2015eggs California (2006)Cormorant South Bay No data No data No data 2,100±1,200 No data Sutton et al. 2015eggs California (2009)Cormorant South Bay No data No data No data 1,100±100 No data Sutton et al. 2015eggs California (2012)Harbor seal San Francisco 1,304 112 87.1 1,730* No data She et al. 2002 (blubber) Bay, California Herring Gull Lake Superior, 253–323 202–284 83.6–113 664–887 (fw)* No data Norstrom et al. Eggs United States (fw) (fw) (fw) 2002 Herring Gull Lake Michigan, 522–602 323–459 167–203 1,366–1,400 No data Norstrom et al. Eggs United States (fw) (fw) (fw) (fw)* 2002 Herring Gull Lake Huron, 146–291 74.6–161 37.3–89.5 308–652 (fw)* No data Norstrom et al. Eggs United States (fw) (fw) (fw) 2002 and CanadaHerring Gull Detroit River, 322 (fw) 130 (fw) 92.6 (fw) 639 (fw)* No data Norstrom et al.Eggs United States 2002Herring Gull Lake Erie, 70–163 52–55.9 24.6–51.8 192–340 (fw)* No data Norstrom et al.Eggs United States (fw) (fw) (fw) 2002Herring Gull Niagara River, 168 (fw) 111 (fw) 53 (fw) 432 (fw)* No data Norstrom et al.Eggs United States 2002Herring Gull Lake Ontario, 220–401 113–322 66.5–102 530–1,003 No data Norstrom et al.Eggs Canada (fw) (fw) (fw) (fw)* 2002Herring Gull St. Lawrence 220 (fw) 89.8 (fw) 56.6 (fw) 453 (fw)* No data Norstrom et al.Eggs River, United 2002StatesHarbor seal St. Lawrence 52–408 No data No data 72–530* No data Fouin et al. 2011Estuary andGulf of St.LawrenceGrey seal St. Lawrence 41 No data No data 69* No data Fouin et al. 2011Estuary andGulf of St.LawrenceHarp seal St. Lawrence 14 No data No data 21* No data Fouin et al. 2011Estuary andGulf of St.LawrenceBeluga Canadian No data No data No data 81–160* No data Alaee et al. 1999whale ArcticBeluga Southeast 10 0.9 1.6 15* No data Stern andwhale Baffin, Canada Ikonomou 2000 PBDEs 4156. POTENTIAL FOR HUMAN EXPOSURETable 6-9. Concentrations (ng/g Lipid Weight) of Several Polybrominated Diphenyl Ethers (PBDEs) in Marine AnimalsSample type Location BDE 47 BDE 99 BDE 100 ΣPBDEsa BDE 209 ReferenceBottlenose South Atlantic No data No data No data 180–220 No data Kuehl et al. 1991dolphin OceanBrunnich’s Svalbard, No data No data No data 130 No data Jansson and guillemot Sweden Asplund 1987Cormorant England, 170– 50–250 50–1,500 300–6,400* No data Allchin et al. 2000United 3,500KingdomCormorant Rhine delta, No data No data No data 28,000 (fw) No data de Boer 1990liver GermanyGalaucous Bear Island, 290–634 160 No data No data No data de Wit 2002gull Norway (Arctic)Grey seal Baltic Sea 650 40 38 730 No data de Wit 2002;Sellström et al.1993Grey seal Baltic Sea 308 54 57 419 No data Haglund et al.1997Grey seal Baltic Sea No data No data No data 208 No data Andersson and Wartanian 1992Harbor British 50– No data No data 350–2,300* No data Ikonomou et al.porpoise Columbia, 1,200 2000CanadaHarbor England and 227– No data No data 440–7,670 No data Law et al. 2000porpoise Wales, United 6,790KingdomHarbor seal Baltic Sea No data No data No data 90 No data Jansson and Asplund 1987Harbor seal Skagerrak, No data No data No data 230 No data Andersson and Norway and Wartanian 1992SwedenHarbor seal North Sea 390– 42–660 25–450 600–6,000 No data de Boer et al.4,900 1998bLong-finned Faeroe Islands 410– 160–600 87–280 843–3,160* No data Lindström et al.pilot whale 1,780 1999Long-finned Faeroe Islands 66–860 24–170 12–98 126–1,250* No data van Bavel et al.pilot whale 1999Minke whale Netherlands 630 160 79 870 No data de Boer et al.1998bRinged seal Baltic sea 256 33 61 350 No data Haglund et al.1997Ringed seal Baltic sea No data No data No data 320 No data Andersson and Wartanian 1992Ringed seal Svalbard, 47 1.7 2.3 51 No data de Wit 2002;Sweden Sellström et al.1993 PBDEs 4166. POTENTIAL FOR HUMAN EXPOSURETable 6-9. Concentrations (ng/g Lipid Weight) of Several Polybrominated Diphenyl Ethers (PBDEs) in Marine AnimalsSample type Location BDE 47 BDE 99 BDE 100 ΣPBDEsa BDE 209 ReferenceRinged seal Canadian ArcticNo data No data No data 25.8–50* No data Alaee et al. 1999Ringed seal Holman Island,Northwest2.8 No data No data 2.4–4.9* No data Ikonomou et al.2000Territories,CanadaSpermwhaleNetherlands 130–250 32–64 21–35 187–349 No data de Boer et al.1998bWhite-beaked Netherlands 5,500 1,000 1,200 7,700 No data de Boer et al.1998bdolphinaΣPBDEs is the sum of BDE 47, BDE 99, and BDE 100, but if more congeners are included, this is marked with an asterisk (*).BDE = brominated diphenyl ether; fw = fresh weight PBDEs 4176. POTENTIAL FOR HUMAN EXPOSUREdifferent sites. In San Francisco Bay, California, the concentrations of PBDEs in harbor seals have increased dramatically based on samples obtained from 1989 to 1998. The samples from 1998 had PBDEconcentrations among the highest reported for this species (She et al. 2002). The concentration of totalPBDEs (sum of BDE 47, BDE 99, BDE 100, BDE 153, and BDE 154) in harbor seal blubber increased by over a factor of 50 from a concentration of 88 ng/g lipid for species samples collected in 1988 to a concentration of 2,985–8,325 ng/g lipid for species samples collected in 1998. The highestconcentrations reported were for BDE 47, which increased from 45.6 ng/g lipid for blubber samples in 1989 to 2,343–6,682 ng/g lipid for blubber samples collected in 1998. The dominance of the tetraBDEcongeners over other congeners may indicate that tetraBDEs bioaccumulate more than the higherbrominated congeners (She et al. 2002). In the Baltic Sea, female grey seals (Halichoerus grypus) sampled in 1979–1985 contained 730 ng PBDE/g lipid in their blubber (sum of BDE 47, BDE 99, and BDE 100) (Jansson et al. 1993); male greyseals had 280 ng PBDE/g lipid weight (Andersson and Wartanian 1992). Male ringed seals (Pusahispida) from the Baltic Sea had 320 ng PBDE/g lipid weight (Andersson and Wartanian 1992). Balticgray and ringed seal blubber sampled between 1981 and 1988 contained 419 and 350 ng PBDEs/g lipid(total of BDE 47, BDE 99, and BDE 100), respectively (Haglund et al. 1997). In 1981, female ringed seals from Svalbard in the Swedish Arctic contained 40–51 ng PBDEs/g lipid in blubber (Jansson and Asplund 1987, Jansson et al. 1993; Sellström et al. 1993). Higher concentrations of PBDEs are generallyevident in Baltic Sea ringed seals (320–350 ng/g lipid) (Andersson and Wartanian 1992; Haglund et al.1997) compared to Arctic ringed seals (26–51 ng/g lipid) (Alaee et al. 1999; Jansson et al. 1987). Theconcentration of PBDEs in harbor seals from Skagerrak on the Swedish west coast was 230 ng PBDE/g lipid (Andersson and Wartanian 1992).She et al. (2000) analyzed the concentration of BDE 47, BDE 99, and BDE 153 in harbor seals from theSan Francisco Bay area. Mean concentrations for BDE 47, BDE 99, and BDE 153 were 1,124, 107, and 50 ng/g lipid weight, respectively. Alaee et al. (1999) found that ringed seals from the Canadian Arctichad mean PBDE concentrations (sum of di- to hexaBDEs) of 25.8 ng/g lipid weight (females) and 50.0 ng/g lipid weight (males). The lower concentrations in female seals suggest that PBDEs are transferred to young through breast milk. On Holman Island, Northwest Territory, Canada (Arctic) in 1996, ringed seals had total PBDE concentrations of 2.4–4.9 ng/g lipid for males. The concentrations ofPBDEs were found to increase with age (Ikonomou et al. 2000). In a temporal trend study, archived samples of blubber from ringed seals from Holman Island, Northwest Territory, Canada were analyzedfor PBDE concentrations. The concentration of PBDE in samples collected between 1981 and 1996 PBDEs 4186. POTENTIAL FOR HUMAN EXPOSUREincreased from approximately 0.3 ng/g lipid weight in 1981 to 3.6 ng/g lipid weight in 1996 (Ikonomou et al. 2000). The concentrations of PBDEs have been determined in harbor porpoises (Phocaena phocaena) from British Columbia, Canada (Ikonomou et al. 2000) and from the coasts of England and Wales (Law et al. 2000). In British Colombia (Canada) samples, the total PBDE concentrations (sum of tri- to heptacongeners) were 350–2,300 ng/g lipid weight; BDE 47 was found at the highest concentrations in these samples (range, 50–1,200 ng/g lipid weight) (Ikononmou et al. 2000). Concentrations of total PBDEs (sum of 13 congeners) along the coast of England and Wales, ranged from 450 to 7,670 ng/g lipid weight, with BDE 47 concentrations ranging from 227 to 6,790 ng/g lipid weight (Law et al. 2000). During a mass mortality event on the south Atlantic coast in 1987–1988, blubber samples were collected from three bottlenose dolphins (Tursiops truncatus); these samples contained 180–220 ng PBDEs/g lipid (Kuehl et al. 1991). Blubber samples, taken from stranded bottlenose dolphins from several locations around the Gulf of Mexico in 1990, contained 3,110 ng PBDEs/g lipid (Kuehl and Haebler 1995). On the Dutch coast in early 1998, de Boer et al. (1998b) found PBDEs in blubber of one whitebeaked dolphin (Lagenorhynchus albirostris); the concentrations of BDE 47, BDE 99, and BDE 100 were 5,500, 1,000, and 1,200 ng/g lipid weight, respectively. The concentration of 19 PBDEs was determined in long-finned pilot whale (Globicephala melas) from the Faeroe Islands in the north Atlantic (Lindström et al. 1999). Young males and females had the highest concentrations, ranging from 3,000 to 3,160 ng/g lipid; lower concentrations were observed for both adult females (840–1,050 ng/g lipid) and males (1,610 ng/g lipid). The predominant isomers in all samples were BDE 47 and BDE 99, accounting for 70% of the sum of 19 congeners. van Bavel et al. (1999) also studied the concentrations of PBDEs in long-finned pilot whales. They observed a similar trend with young animals having higher PBDE concentrations (740 ng/g lipid weight) and adult animals having lower concentrations (females, 230 ng/g lipid; males, 540 ng/g lipid). In Beluga whales sampled in 1997 from southeast Baffin (Cumberland Sound), the concentrations of total PBDEs and BDE 47 were 15 and 10 ng/g lipid weight, respectively (Stern and Ikonomou 2000). Between 1982 and 1997, total PBDE concentrations in archived blubber samples of beluga whales from southeast Baffin Canada increased significantly. For this time period, BDE 47, BDE 99, BDE 100, and BDE 154, and total PBDEs increased by factors of 6.5, 10.3, 7.9, 30.6, and 6.8, respectively (Stern and Ikonomou 2000). Three sperm whales (Physeter macrocephalus) and one minke whale (Balaenoptera acutorostrata) found stranded on the Dutch coast in early 1998 were analyzed for PBDEs (de Boer et al. 1998a, 1998b). Exposure to PBDEs for these animals occurred in the deep Atlantic through the food web. The PBDEs 4196. POTENTIAL FOR HUMAN EXPOSUREconcentrations of PBDEs in these marine mammals were as follows: sperm whale (BDE 47, 130– 250 ng/g lipid weight; BDE 99, 32–64 ng/g lipid weight; and BDE 100, 21–35 ng/g lipid weight) and minke whale (BDE 47, 630 ng/g lipid weight; BDE 99, 160 ng/g lipid weight; BDE 100, 79 ng/g lipidweight); BDE 209 (decaBDE) was below detection limits in all samples.PBDEs, methoxylated (MeO-) PBDEs, and hydroxylated (OH-) PBDEs were evaluated in whole bloodsamples collected from northern fur seal (Callorhinus ursinus), spotted seal (Phoca largha), Steller sealion (Eumetopias jubatus) and ribbon seal (Phoca fasciata) (Nomiyama et al. 2014), and harbor porpoiseand Dall’s porpoise (Ochiai et al. 2013) from northern Japanese coastal waters. The samples contained3.9–280 pg/g median values of 6OH-BDE 47; <1.0–51 pg/g median values of 2’MeO-BDE 68; 2.9– 1,020 pg/g median values of 6MeO-BDE 47; <1–18 pg/g median values of 6MeO-BDE 99; and <100– 230 pg/g median values of total PBDEs.Marine Birds. Increasing concentrations of PBDEs have been found in marine birds and eggs, with BDE 47 found at the highest concentrations. Di- and triBDE have been detected, but not quantified, in black skimmer (Rynchops nigra) tissues and eggs in the United States (Stafford 1983). In 2000, herringgull eggs collected from 15 locations around the Great Lakes (United States and Canada) were pooled and analyzed for PBDEs (Norstrom et al. 2002). A total of 25 di- to hepta-BDE congeners were identified in herring gull through the Great Lakes system. No mono-, octa-, nona-, or decaBDEs were found at the detection limit of the analysis (0.01–0.05 ng/g wet weight). Seven congeners, BDE 28, BDE 47, BDE 99, BDE 100, BDE 153, BDE 154, and BDE 184, constituted 97.5% of total PBDEs (192–1,400 ng/g wetweight). BDE 47 was the dominant congener (70–602 ng/g wet weight) followed by BDE 99 (52– 459 ng/g wet weight). The highest concentrations (1,003–1,400 ng/g wet weight) were found in two LakeMichigan colonies and in Toronto Harbor, Lake Ontario (Norstrom et al. 2002). Muscle tissues fromospreys (Pandion haliaetus), found dead at various locations around Sweden, were pooled and analyzed for PBDEs (Jansson et al. 1993; Sellström et al. 1993). The ospreys’ diet was freshwater fish. Thecombined concentration of BDE 47, BDE 99, and BDE 100 was 2,100 ng/g lipid in samples collectedbetween 1982 and 1986; BDE 47 was the primary congener (86%) in these samples (n=35). Highconcentrations of PBDEs may reflect biomagnification and/or fish consumption along their migratoryroutes. The concentrations of PBDEs in common guillemots (Uria aalge) collected in 1979–1981 fromthe Baltic and North Seas were 370 and 80 ng/g lipid, respectively (Jansson and Asplund 1987). As partof the Swedish National Environmental Monitoring Program, guillemot eggs (St. Karlsö, Baltic Sea) arecollected yearly and placed in the Swedish Natural History Museum’s Environmental Specimens Bank.The concentrations of BDE 47, BDE 99, and BDE 100 in pooled egg samples from the specimen bank PBDEs 4206. POTENTIAL FOR HUMAN EXPOSUREshowed a significant increase from 1969 to the beginning of the 1990s, with highs of 1,100 ng/g forBDE 47 in 1984 and 190 ng/g for BDE 99 in 1990 (Sellström et al. 1993, 1999). Between 1992 and 1997, PBDE concentrations started to decrease statistically. In 1997, the PBDE concentration (sum ofBDE 47, BDE 99, and BDE 100) was 190 ng/g lipid, with BDE 47 as the predominant congener. Cormorant eggs obtained from the San Francisco Bay regions had total PBDE concentrations ranging from 3,425 to 5,550 ng/g (median, 5,500 ng/g) and was dominated by the penta (BDE 47, BDE 99, BDE100) congeners (Klosterhaus et al. 2012). Sutton et al. (20015) noted decreasing levels of PBDEs in cormorant eggs obtained from three locations in northern California. Eggs collected from Suisun, Central, and South Bays in 2012 had total PBDE levels 93, 88, and 74% lower, respectively, when compared to the levels in eggs collected in 2002 (see Table 6-9).Human Body Tissues and Fluids. The quantitative determination of the concentrations of PBDEs in body tissues and fluids is important in determining the human body burden of these chemicals. Increasing concentrations of lower-brominated PBDEs have been measured in blood and breast milk in temporal trend studies. Individuals who consumed fish had a somewhat higher concentration of totalPBDEs in body fluids compared to individuals who ate less fish. Lipid adjusted serum levels of 11 BDE congeners collected from the U.S. general population werereported in the Fourth National Report on Human Exposure to Environmental Chemicals, Updated Tables(CDC 2015). Serum levels for BDE 17, BDE 28, BDE 47, BDE 66, BDE 85, BDE 99, BDE 100, BDE 153, BDE 154, and BDE 183 were evaluated in samples collected between 2003 and 2008; BDE209 was evaluated in samples collected from 2005 to 2008. In the NHANES 2003–2004 survey years, congener BDE 47 was detected at a concentration of 20.5 ng/g lipid (geometric mean), the highestconcentration for all samples. BDE 28, BDE 99, BDE 47, BDE 100, and BDE 153 were in >60% ofparticipants (Sjödin et al. 2008). BDE 17 was not detected above the limit of detection of 1.0 and 0.6 ng/g lipid in survey years 2003–2004 and 2005–2008, respectively. BDE 209 was not detected above thelimit of detection, 6.0 and 5.8 ng/g lipid, in survey years 2005–2006 and 2007–2008, respectively. BDE 47, BDE 99, BDE 100, BDE 153, and BDE 154 were detected in 98, 100, 100, 96, and 48%, respectively, of serum samples collected from 48 mothers participating in the California Childhood Leukemia Study from 2006 to 2007 (Whitehead et al. 2015a). Median serum levels ranged from belowthe detection limit for BDE 154 to 35 ng/g lipid for BDE 47. BDE 47 and BDE 153 were detected in whole blood of 61 and 85%, respectively, of 191 children participating in the California Childhood Leukemia Study from 1999 to 2007 (Whitehead et al. 2015b). The median, 75th percentile, 90th PBDEs 4216. POTENTIAL FOR HUMAN EXPOSUREpercentile, and maximum levels were 410, 820, 1,500, and 17,000 pg/mL, respectively, for BDE 47 and 130, 270, 460, and 6,500 pg/mL, respectively, for BDE 153. Tables 6-10, 6-11, and 6-12 summarize representative concentrations of PBDEs found in blood (serum), adipose tissue, breast milk, and other body tissues or fluids, respectively. These studies indicate that concentrations of lower-brominated BDEs in body fluids are a factor of 10–100-fold higher forindividuals living in the United States compared to individuals living in other regions of the world(e.g., Europe). Serum samples collected from 12 U.S. blood donors in 1988 were analyzed for PBDEs, and BDE 47, BDE 153, BDE 183, and BDE 209 were detected (Patterson et al. 2000; Sjödin et al.2001b). Concentrations of these congeners were similar to those found in the Sjödin et al. (1999b) studyfor the control group. The median concentrations and ranges of BDE 47, BDE 153, BDE 183, BDE 209, and total PBDEs (sum of four congeners) were 0.63 (<0.4–24); 0.35 (0.08–2.0); 0.17 (0.09–1.3); <1 (<1– 34); and 2.2 ng/g lipid weight, respectively (Sjödin et al. 2001b). DecaBDE was found at concentrationsabove the limit of quantification (1 pmol/g lipid) in 5 of 12 serum samples (Patterson et al. 2000). Schecter et al. (2005) provided a summary of PBDE (BDE 17, BDE 28, BDE 47, BDE 66, BDE 77, BDE85, BDE 99, BDE 100, BDE 138, BDE 153, BDE 154, BDE 183, and BDE 209) concentrations in blood from 29 adults residing in Mississippi and 10 adults in New York City. These blood samples wereobtained in 2003. These data were then compared to archived blood samples from 100 individualsobtained in 1973 from the Dallas, Texas area and stored at the University of Texas Southwestern MedicalCenter. The 13 PBDE congeners were not detected in the blood samples from the 100 individualcollected in 1973, but all congeners were detected in at least one of the blood samples obtained from the29 residents of Mississippi collected in 2003. BDE 28, BDE 47, BDE 77, BDE 99, BDE 100, BDE 138, BDE 153, BDE 154, and BDE 183 were all detected in at least one of the blood samples obtained in 2003 from New York City residents. BDE 47 followed by BDE 99 were the predominant congeners detected in both the Mississippi and New York residents blood samples collected in 2003.Rawn et al. (2014) analyzed data from nearly 5,000 serum samples collected from the Canadian HealthMeasures Survey (CHMS) from 2007 to 2009. PBDE congeners were detected in all samples tested, with a range of values of 27–130 ng/g lipid (total PBDEs) and a GM of 46 ng/g lipid. BDE 47 was thepredominant congener with a GM of 22 ng/g lipid followed by BDE 153 (GM=9.4 ng/g lipid), BDE 99 (GM=4.6 ng/g lipid), BDE 100 (GM=4.1 ng/g lipid), and BDE 209 (GM 1.1 ng/g lipid) (Rawn et al. 2014). PBDEs 4226. POTENTIAL FOR HUMAN EXPOSURETable 6-10. Concentrations (ng/g Lipid Weight) of Several PolybrominatedDiphenyl Ethers (PBDEs) in Human Blood SamplesaSample type Location BDE 47 BDE 99 BDE 100 ΣPBDEsa BDE 209 ReferenceHuman United States (in 0.63 0.32 0.17 2.2* <0.1 Sjödin et al. 2001b blood 1988) Human Mississippi and 25 11.1 4.7 (mean) 52.6 2.7 Schecter et al. blood New York City (mean) (mean) (mean) (mean) 2010 (in 2003)Maternal Indiana 28 (9.2– 5.7 (2.4– 4.2 (1.9– 37 (15– No data Mazdai et al. 2003serum 310) 68) 110) 580)Fetal Indiana 25 (8.4– 7.1 (2.2– 4.1 (1.8– 39 (14– No data Mazdai et al. 2003serum 210) 54) 91) 460)Maternal Texas 14.9 3.0 2.8 27.8* Schecter et al.serum 2010Human United States (in 20.5 No data 3.93 (geo- No data No data CDC 2015serum 2003–2004) (geo metricmetric mean)mean)Human United States (in 21.2–74.9 4.16–23.3 4.06–14.2 No data No data CDC 2015serum 2005–2006) (weighted (weighted (weighted arithmetic arithmetic arithmeticmean) mean) mean)Human United States (in 20.5–53.1 4.30–10.7 4.12–11.0 No data No data CDC 2015serum 2007–2008) (weighted (weighted (weighted arithmetic arithmetic arithmeticmean) mean) mean)Human Australia 2.6–55.1 0.9–24.2 0.6–14.1 24 No data Toms et al. 2009serum (newborn);31 (0– 2 years);41 (2– 6 years);26 (7– 12 years);20 (13–30years);9.4 (>31 ye ars) (mean)Maternal France 2.831 1.939 0.365 No data 5.783 Antignac et al.serum 2009Cord France No data 7.434 1.467 No data 27.110 Antignac et al.serum 2009Maternal Spain 2.3 0.35 No data 9.6 <0.7 Vizcaino et al.serum 2011Cord Spain 2.3 1.5 No data 9.6 <1.2 Vizcaino et al.serum 2011 PBDEs 4236. POTENTIAL FOR HUMAN EXPOSURETable 6-10. Concentrations (ng/g Lipid Weight) of Several PolybrominatedDiphenyl Ethers (PBDEs) in Human Blood SamplesaSample type Location BDE 47 BDE 99 BDE 100 ΣPBDEsa BDE 209 ReferenceHuman SerumCanada 23 (mean)5.4 (mean)4.4 (mean) 48 (mean) 1.9 (mean)Rawn et al. 2014Maternal China 0.75 No data No data 29.26 2.12 Li et al. 2013aserum (mean) (mean) (mean)Cord China 0.62 No data No data 41.12 1.33 Li et al. 2013aserum (mean) (mean) (mean)Human Sweden No data No data No data 2.1* No data Klasson Wehler etblood al. 1997Human bloodJapan 0.001 31 year olds were 24, 31, 41, 26, 20, and 9.4 ng/g lipid, respectively. The peak mean concentration was observed in toddlers 2.6– 3 years of age (51 ng/g lipid), which is later than when breastfeeding usually ceases, suggesting a lowercapacity to eliminate PBDEs or greater exposure through unique exposure pathways more common forthis age group (e.g., ingestion or dermal exposure of contaminated dust particles in carpeting).BDE 47, BDE 153, BDE 154, BDE 183, and BDE 209 were measurable in blood plasma from three groups of workers (i.e., workers at a computer-disassembly plant, workers in a computerized office, and acontrol group) (Sjödin et al. 1999a). The median concentrations (sum of five congeners) were highest forthe computer-disassembly plant workers (26 ng/g lipid weight); the office workers had a median concentration of 4.1 ng/g lipid weight and the control group had a median concentration of 3.3 ng/g lipidweight. The congener patterns for the control group and office workers were similar, with BDE 47 having the highest concentrations. For the computer disassembly plant workers, the median concentrations of BDE 183, BDE 153, BDE 154, BDE 47, and BDE 209 were 7.8, 4.5, 1.2, 2.9, and PBDEs 4286. POTENTIAL FOR HUMAN EXPOSURE4.8 ng/g lipid weight, respectively. Blood serum samples from 19 full-time computer technicians were analyzed (Hagmar et al. 2000a). The serum concentrations of BDE 153, BDE 183, and BDE 209 in thesesamples were found to be approximately 5 times higher than the control and office workers in the Sjödin et al. (1999a) study. The median concentration for total PBDEs (for the sum of five congeners) was10.6 pmol/g (7.0 ng/g) lipid weight. The highest concentrations were of BDE 153. Two octaBDEcongeners and one nonaBDE congener were also detected. Connections were observed between fish consumption and serum concentrations for congeners BDE 47, BDE 153, and BDE 183, and between worktime at the computer and congeners BDE 153 and BDE 183.DecaBDE, as well as hexa- through nonaBDE, has been found in composite samples from the 1987 National Human Adipose Tissue Survey repository (Cramer at al 1990; Stanley et al. 1991). The concentrations ranged from not detected to 1 ng/g fat for hexaBDE, 0.001–2 ng/g fat for heptaBDE andnot detected to 8 ng/g fat for octaBDE. NonaBDE concentrations were estimated to be >1 ng/g fat;decaBDE was estimated to range between not detected and 0.7 ng/g fat. In the late 1990s, breast adiposesamples collected in northern California contained quantifiable amounts of BDE 47, BDE 99, and BDE 153 (She et al. 2000). Mean concentrations were 18 ng/g lipid weight for BDE 47, 4.9 ng/g lipidweight for BDE 99, and 2.2 ng/g lipid weight for BDE 153. Average total PBDEs concentrations(86 ng/g lipid) were the highest human concentrations reported to date. Petreas et al. (2003) expandedtheir investigation to include a diverse group of local women from the San Francisco Bay area sampled in the late 1990s. Their results confirmed earlier findings reported in She et al. (2000). Mean concentrations of BDE 47 in adipose tissues samples taken from California women were 28.9 ng/g lipid.In the adipose tissue of a 74-year-old Swedish male, the BDE 47 concentration was 8.8 ng/g lipid weight (Haglund et al. 1997). Adipose and liver tissue from two Swedish males were examined for several PBDEs (BDE 28, BDE 47, BDE 85, BDE 99, BDE 100, BDE 153, and BDE 154) (Meironyté Guvenius and Norén 1999). Thedistribution of congener concentrations in the adipose and liver tissues for each individual were similar.BDE 47, BDE 99, and BDE 153 were the predominant congeners with adipose BDE 47 concentrationsranging from 2 to 2.4 ng/g lipid weight, BDE 99 concentrations of 1.6 ng/g lipid weight, BDE 100 concentrations of 0.1 ng/g lipid weight, and BDE 153 concentrations ranging from 1 to 1.3 ng/g lipidweight. The total PBDE concentration (i.e., the sum of the seven congeners) in adipose tissue was 5 ng/g lipid weight. Human liver and adipose tissues from one woman and four men autopsied in Sweden in 1994 were analyzed for PBDEs containing 3–6 bromine atoms (Meironyté Guvenius and Norén 2001).PBDEs were found in all of the tissue samples. The sums of nine congeners (BDE 17, BDE 28, BDE 47, PBDEs 4296. POTENTIAL FOR HUMAN EXPOSUREBDE 66, BDE 100, BDE 99, BDE 85, BDE 154, and BDE 153) were 5–18 and 4–8 ng/g lipids in liverand adipose tissue, respectively. The PBDE congeners BDE 47, BDE 99, and BDE 153 occurred at thehighest concentrations and constituted 87–96 and 84–94% of the total sum in liver and adipose tissue, respectively. Strandman et al. (1999) measured the concentration of BDE 47, BDE 99, and BDE 153 in adipose tissue samples from 10 randomly selected individuals in Finland. Mean concentrations were 7.3 ng/g fat for BDE 47, 2.2 ng/g fat for BDE 99, and 2.3 ng/g fat for BDE 153. Concentrations ofPBDEs were measured in adipose tissue samples from 13 individuals (3 women, 10 men) fromTarragona, Spain; the mean concentrations of BDE 47, BDE 99, and BDE 153 were 1.36, 0.42, and1.83 ng/g lipid weight, respectively. The mean concentrations of pentaBDE and hexaBDE were 0.93 and 1.83 ng/g lipid weight, respectively (Meneses et al. 1999).Human Milk. Schecter et al. (2003) reported the first findings on concentrations of PBDEs congeners in human milk from individuals in the United States. Forty-seven individual milk samples were analyzedfrom nursing mothers, 20–41 years age, from a milk bank in Austin, Texas, and a community health clinic in Dallas, Texas, both in the year 2001. The median concentration of the sum of PBDE congeners was34.0 ng/g lipid. The predominant congener was BDE 47 (18.4 ng/g lipid); other congeners detected were BDE 17, BDE 28, BDE 66, 2, BDE 85, BDE 99, BDE 100, BDE 138, BDE 153, BDE 154, and BDE 183 at median concentrations of 0.01, 1.2, 0.14, 0.41, 5.7, 2.9, 0.09, 2.0, 0.22, and 0.07 ng/g lipid, respectively. DecaBDE was detected in 7 out of 47 samples with a maximum concentration of 8.24 ng/g lipid. PBDE concentrations in breast milk from this study were similar to concentrations found in U.S. blood and adipose tissue lipid from California and Indiana and are 10–100 times greater than human tissue concentrations in Europe (Schecter et al. 2003). These data have been updated to include 2001– 2004 samples and are provided in Table 6-13 (Schecter et al. 2005). Median concentrations of BDE 47, BDE 100, and BDE 153 were 26, 4, and 3.5 ng/g lipid, respectively, in breast milk samples collected from 10 mothers in Pennsylvania (LaKind et al. 2009). The detection frequency was 100% for each congener in all 35 samples collected.Norén and Meironyté (1998, 2000) examined the temporal trends of PBDE concentrations in pooled breast milk samples from mothers in Stockholm, Sweden. Between 1972 and 1997, the concentration ofPBDEs in human breast milk increased, with a doubling rate of 5 years. In the 1997 sample, theconcentration of PBDEs (sum of eight congeners) was 4 ng/g lipid, whereas the 1972 sample contained 0.07 ng/g lipids (Meironyté et al. 1999). The authors suggest that the current exposure of humans toPBDEs may not be only diet; other exposure routes may result from the presence of PBDE in both work PBDEs 4306. POTENTIAL FOR HUMAN EXPOSURETable 6-13. Concentrations of PBDE Congeners in Human Milk from Nursing Mothers in the United States (2001–2004) (ng/g Lipid)BDE congenerFemaleLipid(%)Age(years)Nursing (weeks) 17 28 47 66 77 85 99 100 138 153 154 183 209 Sum1a 4.8 31 3 ND 0.2 2.9 0.02 ND 0.08 0.7 0.7 ND 1.5 0.06 ND ND 6.22a 1.3 29 3 ND 0.3 3.5 ND ND 0.08 0.7 0.5 ND 0.9 0.06 0.06 ND 6.23a 2.1 23 74 ND 0.2 3.9 0.06 ND 0.11 1.5 0.6 ND 0.4 0.08 ND ND 6.94b 4.8 32 21 — 0.3 3.5 0.14 — 0.08 1.6 0.7 0.09 1.4 0.09 0.04 — 85a 2.6 22 40 0.01 0.3 6.3 0.05 ND 0.23 2.8 1.2 ND 0.7 0.2 0.05 ND 11.86a 3.6 36 109 ND 0.4 7.8 0.09 ND 0.23 2.4 1.1 0.01 0.4 0.11 ND ND 12.57a 1.9 32 20 ND 0.7 8.2 0.04 ND 0.22 1.3 1.7 ND 0.9 0.12 0.08 ND 13.38a 6.3 25 2 ND 0.4 7.9 0.02 ND 0.38 2.3 2.7 ND 0.8 0.15 0.06 ND 14.79b 2.1 35 29 — 0.7 8.8 0.19 — 0.17 1.5 1.7 0.16 2 0.06 0.03 — 15.210a 5.5 32 30 ND 0.4 8 0.01 ND 0.44 2.9 2 ND 0.9 0.14 0.24 0.48 15.611b 5 20 2 0.01 1.1 10.9 0.05 ND 0.18 2 2.4 ND 1.3 0.17 0.04 ND 18.112a 3.4 23 3 0.01 0.4 8 0.03 ND 0.35 3.1 2.7 ND 2 0.21 0.61 0.93 18.313b 1.3 32 16 — 0.8 10.5 ND — 0.35 2.5 2.2 0.19 2 0.12 0.03 — 18.614a 3.4 25 NA 0.02 1.1 12 0.13 ND 0.23 2.5 1.8 ND 1.3 0.15 0.08 1.85 21.115a 2.9 21 29 0.03 0.5 10.7 0.09 ND 0.27 5.5 2.1 ND 0.9 0.35 0.07 2.74 22.416b 3.5 30 30 — 0.7 6.9 ND — 0.12 1.3 4.6 0.41 8.5 0.19 0.06 — 22.817a 1 23 2 ND 0.9 14.2 0.11 ND 0.37 3.7 2.6 ND 1.3 0.24 0.09 ND 23.518b 3.7 23 19 — 0.7 13.2 0.57 — 0.29 3.7 2.5 0.29 1.9 0.17 0.06 — 23.519a 3.2 26 2 ND 1.3 17.4 0.19 ND 0.35 4 2.1 ND 0.7 0.18 ND ND 26.220b 3.5 34 22 — 1 14.3 0.29 — 0.46 5.7 3.6 0.25 1.4 0.2 0.1 — 27.321b 3.1 33 60 — 1.4 18.4 ND — 0.25 4.1 1.8 0.09 2.1 0.16 0.06 — 28.322b 4.9 38 26 — 1.2 17.4 ND — 0.34 7.1 2.3 0.14 0.6 0.3 0.12 — 29.623a 3.4 30 2 0.01 0.7 15.2 0.06 ND 0.42 4.2 2.3 ND 3 0.22 0.03 3.97 30.124a 5.1 28 53 0.01 1.1 20 0.18 ND 0.53 5.1 3.9 0.01 2.7 0.32 0.11 ND 3425b 4.7 35 NA — 1.3 20.9 0.56 — 0.31 6.3 2.9 0.14 1.2 0.22 0.17 — 34.126a 1.1 41 38 0.02 1.5 19.5 0.11 ND 0.41 3.4 3.3 ND 7.7 0.18 ND ND 36.127b 6.1 37 25 — 7.6 17.2 1.19 — 0.35 6.1 2.3 0.18 1.7 0.27 0.05 — 36.828b 3 27 51 — 1.4 28.2 ND — 0.51 7.5 2.9 0.25 0.7 0.2 0.75 — 42.429b 4.8 25 NA — 1.8 21.6 0.94 — 0.5 9.4 4.4 0.47 5.8 0.6 0.06 — 45.530b 2.2 39 11 — 1.1 26.8 ND — 0.75 8.9 5.3 0.58 2 0.45 0.1 — 4631b 5.6 34 NA — 2.7 31.8 ND — 0.42 7.8 3.1 0.09 0.8 0.22 0.1 — 4732b 3.4 27 10 — 2.6 30.1 0.75 — 0.57 5.9 6.5 0.32 2.5 0.34 0.09 — 49.633a 2.8 20 13 0.02 1.1 31.3 0.17 ND 0.13 10.2 5.9 0.02 1.5 0.48 0.11 2.96 53.934b 4 20 13 — 3.4 33.5 2.32 — 0.49 5.8 5.8 0.27 2.6 0.32 0.06 — 54.635b 3.3 26 17 — 1.4 32.3 0.7 — 0.66 9.6 5.7 0.46 12.4 0.51 0.05 — 63.836b 2.2 20 16 — 1.4 25.5 0.75 — 0.76 8 18.3 1.75 18.3 0.94 0.08 — 75.837a 1.1 22 51 0.04 2.2 44.3 0.55 0.03 0.64 10.8 8.1 0.02 14.7 0.56 ND ND 81.938b 4.3 29 38 — 5.2 34.8 0.61 — 1.94 9.8 29.2 1.2 14.5 0.96 0.1 — 98.2 PBDEs 4316. POTENTIAL FOR HUMAN EXPOSURETable 6-13. Concentrations of PBDE Congeners in Human Milk from Nursing Mothers in the United States (2001–2004) (ng/g Lipid)BDE congenerFemaleLipid(%)Age(years)Nursing (weeks) 17 28 47 66 77 85 99 100 138 153 154 183 209 Sum39a 1 26 22 0.02 1.7 54.7 0.54 ND 1.63 23.6 10 ND 4.8 1.15 0.07 ND 98.240b 4.9 32 38 — 3.4 49.7 1.21 — 1.2 7.7 21.1 1.4 16.3 0.93 0.15 — 103.141b 3.4 30 9 — 3.9 63.1 3.13 — 2.81 30.1 16.2 3.29 17.2 1.94 0.08 — 141.642b 1.2 21 2 0.1 10.1 120.9 1.68 0.06 2.64 30.3 20.1 0.13 16.4 2.07 ND ND 204.343b 1.2 33 15 — 8 139.6 ND — 4.12 44.6 23 4.47 21.8 2.76 0.18 — 248.544a 1 23 2 0.06 3.6 172.4 1.14 ND 6.28 69.8 31.9 0.08 8.4 3.07 0.16 ND 296.945b 2.1 34 13 — 7.5 199.6 6.67 — 7.73 108.5 31.7 4.12 6.9 3.62 0.36 — 376.746a 1.7 33 47 0.18 6.1 196.2 2.07 0.16 6.46 111 31 0.27 15.5 7.21 1.32 8.24 385.547b 5.1 29 28 — 16.1 271.5 3.16 — 6.29 50.4 47.4 6.86 14.1 2.87 0.12 — 418.848 3.18 NA 0.011 0.6 9.9 0.058 NA NA 2.7 1.3 0.022 0.62 0.13 0.072 2.4 17.8149 3.28 NA 0.016 1.2 25 0.12 NA NA 11 4.5 0.092 3.1 0.57 0.092 0 45.6950 3.9 NA 0.016 4.4 41 0.098 NA 1.3 11 11 0.15 6.1 0.71 0.14 0 75.9151 3.12 NA 0.015 2.3 37 0.11 NA 0.49 6.7 6.9 0.086 11 0.39 0.11 0 65.1052 6.31 NA 0.01 1.8 14 0.084 NA NA 2.9 2.9 0.033 2.1 0.17 0.1 0 24.153 7.04 37 0.02 0.71 16 0.13 NA 0.46 5.8 2.5 0.03 2.2 0.28 0.057 0.02 28.2154 3.29 33 0.014 0.76 9 0.056 NA 0.14 1.5 2.1 0.01 2.6 0.094 0.051 0.05 16.3755 2.44 38 0.02 1.2 19 0.18 NA 0.35 5 2.7 0.02 2.4 0.27 0.051 2.5 33.6956 6.5 NA 0.005 0.9 6.1 0.088 NA 0.082 1 1.2 0.01 6.7 0.077 0.029 0.03 16.2257 5.55 NA 0.005 0.9 6.1 0.088 NA 0.082 1 1.2 0.01 3.4 0.077 0.029 0.03 12.9258 6.75 NA 0.046 1.5 24 0.18 NA 0.38 4.8 2.7 0.02 2.1 0.22 0.57 0.05 36.5759 3.08 NA 0.016 1.2 10 0.086 NA 0.18 2 1.3 0.01 4 0.15 0.096 0.1 19.14aAustin milk bank sample.bDallas milk bank sample.NA = not available; ND = not detectedSource: Schecter et al. 2005 PBDEs 4326. POTENTIAL FOR HUMAN EXPOSUREand home environments. PBDE concentrations were studied in breast milk obtained from motherspregnant for the first time (n=39, ages 22–36 years old) from Uppsala County, Sweden (Darnerud et al. 1998). The mean value of total PBDEs (sum of eight congeners) was 4.4 ng/g fat; the major congenerwas BDE 47, contains ca. 55% of the total PBDEs. Lind et al. (2003) reported concentrations of PBDEsin human breast milk sampled from Uppsala County, Sweden. Total PBDEs, BDE 47, BDE 99, andBDE 100 concentrations were 4.01, 2.35, 0.62, and 0.38 ng/g lipid, respectively. In human breast milkfrom 25 German mothers, the concentrations of PBDEs ranged from 0.6 to 11 ng/g lipid (de Wit 2002). In 1992, the mean concentration of total PBDEs (sum of BDE 28, BDE 47, BDE 99, BDE 100, BDE 153, and BDE 183) was 5.8 ng/g lipid weight for samples (n=6) from mothers from Ontario and Quebec, Canada (Ryan and Patry 2000). Combined samples from 1992 representing four regions of Canada and one representing all Canadian provinces had total PBDE concentrations ranging from 2.6 to 19 ng/g lipid weight; the highest concentrations were observed in the New Brunswick, Nova Scotia, and Prince Edward Island. Breast milk samples from Finland, collected between 1994 and 1998, had concentrations of totalPBDEs (sum of BDE 28, BDE 47, BDE 99, and BDE 153) ranging from 0.88 to 5.9 ng/g lipid weight (Strandman et al. 2000). In Japan, breast milk samples had total PBDE concentrations (sum of BDE 28, BDE 47, BDE 99, BDE 100, BDE 153, and BDE 154) ranging from 0.66 to 2.8 ng/g lipid weight (Ohta et al. 2002). Women who consumed fish had a somewhat higher concentration of total PBDEs (range, 1.4– 2.8 ng/g lipid weight) compared to women who ate less fish (range, 0.67–0.87 ng/g lipid weight).BDE 47 was the major congener in most of the samples; BDE 153 concentrations were analogous to BDE 47 concentrations in some samples (Ohta et al. 2002). Hydroxy- and Methoxy- Derivatives in Biota. Hydroxy- and methoxy- derivatives of PBDEs have beenidentified in biota. However, their origin in the environment has not yet been explained. Anthropogenicsources of these compounds have not been found. Tetra- and pentabrominated methoxy (MeO) BDEswere found in herring, salmon, grey seal, ringed seal, and white-tailed sea eagle from the Baltic region(Asplund et al. 1999a; Haglund et al. 1997) as well as beluga whale from Svalbard and pilot whale fromthe Faroe Islands (van Bavel et al. 2001). The concentrations of hydroxy- and methoxy- derivatives were of the same order of magnitude as PBDEs present in the samples. Biogenic production via metabolism ofPBDEs or natural production via biobromination has been suggested as the origin for these compounds. Naturally produced methoxy-tetrabrominated diphenyl ethers have been reported in tropical marine sponges (sp. Dysidea) as well as in green algae (sp. Cladophora) collected in Japan (Kierkegaard et al.2004). Kierkegaard et al. (2004) found that the concentrations of 6-methoxy-2,2’,4,4’-tetraBDE inherring from five locations along the Swedish coast increased from south to north in the Baltic Sea. No PBDEs 4336. POTENTIAL FOR HUMAN EXPOSUREcorrelation between the concentrations of BDE congeners and methoxy-brominated diphenyl ethers wasobserved, indicating sources other than PBDEs for these compounds. Biomonitoring Historical Trends and Future Projections. Concentrations of PBDE in human and animal tissues have increased since their development and widespread use as flame retardants incommercial products. Blood samples collected from U.S. residents in 1973 did not contain measurable concentrations of PBDE congeners; however, many congeners have been identified at varyingconcentrations in U.S. blood samples since the widespread use of PBDEs as flame retardants (Schecter etal. 2005). For example, the CDC reported BDE congener concentration in serum collected in 2003–2004from the general U.S. population (CDC 2015). The total geometric mean concentrations ranged frombelow the limit of detection for BDE 17 to 20.5 ng/g lipid for BDE 47. In general, body burden concentrations of PBDEs in North America are higher than in Europe due to higher historical demand and usage. Since all production and use of penta-, octa-, and now decaBDE have ceased in the United States, future biomonitoring results will likely show a gradual decline in body burden concentrations of thesesubstances in U.S. residents as products containing PBDEs ultimately become rare. Age-dependent dataon PBDE levels indicate several sources of human exposure. Serum samples collected in Australiasuggest that PBDE levels increase from infant to toddler and then gradually decrease over time (Toms etal. 2009). Peak average concentrations were observed in toddlers 2.6–3 years of age (51 ng/g lipid),which is later than when breastfeeding usually ceases, suggesting a lower capacity to eliminate PBDEs orgreater exposure through unique exposure pathways more common for this age group (e.g., ingestion ordermal exposure of contaminated dust particles in carpeting). The EPA Exposure Assessment ofPolybrominated Diphenyl Ethers published in May of 2010, summarizes many other biomonitoringstudies not discussed here and the reader is encouraged to consult this assessment for additional analysis of the environmental fate and biomonitoring of PBDEs (EPA 2010).6.5 GENERAL POPULATION AND OCCUPATIONAL EXPOSUREHumans are exposed to PBDEs by a wide variety of routes including ingestion of contaminated foodsinhalation of air, ingestion of contaminated dusts/soils, and dermal exposure routes. The EPA published an exposure assessment of the U.S. population to PBDEs and determined that the overall weight-ofevidence suggested that bulk of U.S. exposures occur in indoor environments through ingestion and contact with house dust. It concluded that house dust accounts for between 80 and 90% of total exposuresof the general population, with the remainder due primarily to food ingestion. Watkins et al. (2011) PBDEs 4346. POTENTIAL FOR HUMAN EXPOSUREdetermined that regular hand washing decreases the mass of PBDEs on hands from dust samples and is thus expected to reduce intake from hand-to-mouth activities.Intake doses of BDE 47, BDE 99, BDE 100, and BDE 153 from all exposure pathways for the North American population were modeled by Wong et al. (2013). The model assumed intake of PBDEsincreased exponentially to a peak in 2004, and has since exponentially declined. The intakes of BDE 47, BDE 99, BDE 100, and BDE 153 were estimated as 3.88–54, 1.59–2.39, 1.17–2.98, and 1.37–2.44 ng/kgbody weight/day depending upon how the intakes were fit to measured body burden data using differentelimination half-lives. Trudel et al. (2011) used eight different exposure pathways (oral uptake of food, dust, soil, and organic films; inhalation of air; and dermal uptake of dust, soil, and organic films) to modelintakes of PBDEs for different age/gender groups in North American and European populations. Themean intakes for total PBDEs in the North American population were 210.0, 80.0, 79.0, 69.0, 43.0, 28.0, and 22.0 ng/kg body weight/day for infants, toddlers, children, female teenagers, male teenagers, female adults, and male adults, respectively. These concentrations are about 3–8 times greater than the estimatedintakes for European populations. Lorber (2008) also estimated PBDE intake of the U.S. populationthrough similar exposure routes. The adult intake of total PBDEs was estimated as 7.7 ng/kg bodyweight/day, while the intake of children aged 1–5 years was 49.3 ng/kg body weight/day. The intakes for6–11 year olds and 12–19 year olds were estimated as 14.4 and 9.1 ng/kg body weight/day, respectively(Lorber 2008). Exposure from indoor house dust accounted for about 82% of the intake (66% fromsoil/dust ingestion, 16% from soil/dust dermal contact) of total PBDEs, while inhalation and ingestion offood and water accounted for <20% of the total intake. BDE 47, BDE 99, BDE 100, and BDE 209 werethe predominant congeners, accounting for 26, 28, 11, and 27%, respectively, of the total intake (Lorber2008). The EPA 2010 Exposure Assessment of Polybrominated Diphenyl Ethers calculated the adultintake dose of total PBDEs to be 7.1 ng/kg body weight/day (EPA 2010). The largest source contributingto PBDE exposure in the United States was reported to be house dust (ingestion and dermal exposure),contributing about 90% of the overall estimated intakes. The EPA exposure assessment estimatedchildren intakes as 47.2 ng/kg body weight/day for 1–5 year olds, 13.0 ng/kg body weight/day for 6– 11 year olds, and 8.3 ng/kg body weight/day for 12–19 year olds. Intake modeling using a breastfeedingpathway, which used measured milk concentrations and infant ingestion rates of human milk, led to estimated infant intakes of 141 ng/kg body weight/day (EPA 2010). While exposure to dust appears to bethe predominant exposure pathway for the general population of North American residents, PBDEexposure through dietary routes appears to be more important for European communities (Abdallah and Harrad 2014; Law et al. 2008). PBDEs 4356. POTENTIAL FOR HUMAN EXPOSUREBreast adipose samples collected in northern California in the late 1990s contained quantifiable amountsof BDE 47, BDE 99, and BDE 153 (She et al. 2000). Mean concentrations were 18 ng/g lipid weight forBDE 47, 4.9 ng/g lipid weight for BDE 99, and 2.2 ng/g lipid weight for BDE 153. In studies of thegeneral populations of other countries, it has also been shown that exposure to lower-brominated PBDEcongeners by the general population is widespread (see Section 6.4.4; Haglund et al. 1997; Meneses et al.1999). In general, concentrations of decaBDE in human tissues and body fluids are lower than for thelower-brominated congeners, presumably due to a more rapid elimination half-life (Trudel et al. 2011). Consumption of fish has been associated with elevated concentrations of PBDEs in tissues from theSwedish population (Bergman et al. 1999). In Sweden, fish consumption is about 30 g/day; this translates to an estimated 0.1 μg of pentaBDE and 0.3 μg of total PBDEs from fish that is ingested by humans daily(WHO 1994a). The fish of greatest concern to humans are bottom feeders like carp and catfish. Harradet al. (2004) estimated the daily dietary intakes of PBDEs in omnivorous and vegan diet samples from theUnited Kingdom. In this study, the median lower bound estimates of dietary exposure (i.e., where acongener is below the detection limit, the concentration is assumed to be zero) for BDE 47, BDE 99, BDE 100, BDE 153, BDE 154, and total PBDEs were 46.4, 42.6, 0, 0, 0, and 90.5 ng/day, respectively(Harrad et al. 2004). The International Polar Year Inuit Health Survey in 2007–2008 evaluated PBDEblood concentration for 2,172 Inuit adults in Canada (Laird et al. 2013). The sum concentration ofBDE 47, BDE 99, and BDE 100 in the general population ranged from 0.04 to 10.6 μg/L in blood plasma. Like PCBs, there may be a higher risk of exposure to PBDEs in Native Americans who reside in the Arctic region and consume whale and seal blubber (Jaret 2000).Workers involved in the production and manufacture of PBDE-containing plastics and plastic productsare exposed to PBDEs. Body burden data indicate higher concentrations for workers exposed to PBDEsthan for the general population. Occupational exposure to PBDEs also occurs in workers at plants thatdismantle electronic equipment, computer monitor repair technicians, and automobile drivers, as well asother professions (Lindström 1999). Occupational exposure occurs primarily by inhalation. Inhalation ofvapor-phase PBDEs is expected to be low due to the low vapor pressures of PBDEs (see Table 4-4);however, the inhalation of particulate phase PBDEs is possible during plastic reprocessing where grindingor shredding of polymers with PBDEs occurs. Occupational exposure may also likely involve oralexposure to particulate PBDEs as a result of hand-to-mouth activity.Air samples were taken from an electronics dismantling plant, an office with computers, and outdoors and then analyzed for PBDEs (Sjödin et al. 1999a, 2001a). The electronics dismantling plant had the highest PBDEs 4366. POTENTIAL FOR HUMAN EXPOSUREconcentrations of PBDEs, with mean concentrations of 2.5 pmol/m3 (1.25 ng/m3) for BDE 47, 4.6 pmol/m3 (2.6 ng/m3) for BDE 99, 6.1 pmol/m3 (3.93 ng/m3) for BDE 153, 26 pmol/m3 (18.8 ng/m3) for BDE 183, and 38 pmol/m3 (36.5 ng/m3) for decaBDE (BDE 209) (Sjödin et al. 1999a, 2001a). Airsamples were found to be 4–10 times higher in PBDE concentrations near a plastic shredder whencompared to other locations in the plant (range, 0.42–200 ng/m3). Concentrations of PBDEs in the office(range, <0.002–0.09 ng/m3) were 400–4,000 times lower than in the plant, and PBDEs were not detectedin outside air (Sjödin et al. 1999a, 2001a). Lipid adjusted serum levels of 11 BDE congeners collected from the US general population were reported in the Fourth National Report on Human Exposure to Environmental Chemicals, Updated Tables (CDC2015; also see http://www.cdc.gov/biomonitoring/ ). Serum levels for BDE 17, BDE 28, BDE 47, BDE 66, BDE 85, BDE 99, BDE 100, BDE 153, BDE 154, and BDE 183 were evaluated in samplescollected between 2003 and 2008; BDE 209 was evaluated in samples collected from 2005 to 2008. In the NHANES 2003–2004 survey years, congener BDE 47 was detected at a concentration of 20.5 ng/glipid (geometric mean), the highest concentration for all samples. BDE-153 had the second highestgeometric mean concentration of 5.7 ng/g lipid. BDE 28, BDE 99, BDE 47, BDE 100, and BDE 153 were in ≥60% of participants (Sjödin et al. 2008). The serum levels of BDE 47, BDE 99, and BDE 153 were highest in the youngest age group (12–19 years old) and decreased for the older age groups (from20–39 to 40–59 years old) and then increased in the ≥60 years old age group. BDE 47, BDE 153, BDE 154, BDE 183, and decaBDE (BDE 209) were measurable in blood plasma fromthree groups of workers (i.e., workers at a computer disassembly plant, workers in a computerized office, and a control group) (Sjödin et al. 1999b). The median concentrations (sum of five congeners) were highest for the computer disassembly plant workers (26 ng/g lipid weight); the office workers had amedian concentration of 4.1 ng/g lipid weight and the control group had a median concentration of3.3 ng/g lipid weight. The congener patterns for the control group and office workers were similar, with BDE 47 having the highest concentrations. For the computer disassembly plant workers, the median concentrations of BDE 183, BDE 153, BDE 154, BDE 47, and BDE 209 were 7.8, 4.5, 1.2, 2.9, and 4.8 ng/g lipid weight, respectively. Blood serum samples from 19 full-time computer technicians were analyzed (Hagmar et al. 2000a). The serum concentrations of BDE 153, BDE 183, and BDE 209 in thesesamples were found to be approximately 5 times higher than the control and office workers in the Sjödin et al. (1999b) study. The median concentration for total PBDEs (for the sum of five congeners) was10.6 pmol/g (7.0 ng/g) lipid weight. The highest concentrations were for BDE 153. Two octaBDEcongeners and one nonaBDE congener were also detected. PBDEs 4376. POTENTIAL FOR HUMAN EXPOSURE6.6 EXPOSURES OF CHILDREN This section focuses on exposures from conception to maturity at 18 years in humans. Differences fromadults in susceptibility to hazardous substances are discussed in Section 3.7, Children’s Susceptibility.Children are not small adults. A child’s exposure may differ from an adult’s exposure in many ways. Children drink more fluids, eat more food, breathe more air per kilogram of body weight, and have a larger skin surface in proportion to their body volume. A child’s diet often differs from that of adults. The developing human’s source of nutrition changes with age: from placental nourishment to breast milkor formula to the diet of older children who eat more of certain types of foods than adults. A child’sbehavior and lifestyle also influence exposure. Children crawl on the floor, put things in their mouths, sometimes eat inappropriate things (such as dirt or paint chips), and spend more time outdoors. Children also are closer to the ground, and they do not use the judgment of adults to avoid hazards (NRC 1993).Body burden data, as well as intake modeling, suggest that infants and toddlers have higher exposures toPBDEs as compared to older children or adults. PBDE concentrations increase from infant to toddler and then gradually decrease over time. PBDE intake values for children have been estimated using severalmodels (EPA 2010; Lorber 2008; Trudel et al. 2011; Wong et al. 2013). Using the model developed bythe EPA, total PBDE intakes for children residing in the United States were estimated as 47.2 ng/kg bodyweight/day for 1–5 year olds, 13.0 ng/kg body weight/day for 6–11 year olds, and 8.3 ng/kg bodyweight/day for 12–19 year olds. Data from fetal tissue and several studies, including measurements ofPBDE congeners from umbilical cord blood, indicate that the fetus is exposed to PBDEs through themother. Schecter et al. (2003, 2005) reported the first findings on concentrations of PBDEs congeners in human milk from individuals in the United States. The median concentration of the sum of PBDE congeners was34.0 ng/g lipid with BDE 47 (18.4 ng/g lipid) as the predominant congener. DecaBDE was detected in 7 out of 47 samples with a maximum concentration of 8.24 ng/g lipid. The concentrations of PBDEs in breast milk from this study were 10–100 times greater than human tissue concentrations in Europe (Schecter et al. 2003). Guo et al. (2016) found that PBDE levels in breast milk samples declined in females from California over two sampling periods (2003–2005 and 2009–2012). The geometric mean oftotal PBDE congeners (sum of BDE 28, 47, 99, 100, 153, and 154) in breast milk was 67.8 ng/g lipid in the 2003–2005 sampling period (n=82) and 45.7 ng/g lipid in the 2009–2012 sampling period (n=66). PBDEs 4386. POTENTIAL FOR HUMAN EXPOSUREPBDE levels were shown to increase in human milk samples collected from different regions of Canada from 1992 to 2002, but declined slightly from 2002 to 2005 (Ryan and Rawn 2014). The median and geometric mean levels of total PBDEs (sum of BDE 28, 47, 99, 100, 153, 154, and 183) in collected milksamples were 2.992 and 3.536 ng/g lipids in 1992, respectively, and 22.104 and 25.162 ng/g lipid, respectively, in 2002. The median and geometric mean of total PBDEs in milk samples collected inMarch and April 2005 were 19.948 and 21.082 ng/g lipid, respectively. PBDEs were detected in human placental tissues (n=102) collected between 2010 and 2011 in DurhamCounty, North Carolina (Leonetti et al. 2016). The geometric mean concentration of total PBDE (sum ofBDE 47, 99, 100, 153, 154, and 209) was 17.6 ng/g lipid. The detection frequencies of the individualcongeners were: BDE 47, 91.2%; BDE 99, 68.6%; BDE 100, 88.2%, BDE 153, 93.1%; BDE 154, 83.3%; and BDE 209, 52.9%. PBDEs were detected at a median concentration of 53,000 ng/g in a variety of toys such as hard plastictoys (racing cars, vehicles, toy weapons, etc.), foam toys, rubber/soft plastic toys (dolls and teethers), andstuffed toys that were purchased in China from 2007 to 2008 (Chen et al. 2009). These findings suggestadditional possible exposure routes to children and toddlers through mouthing activities and dermalcontact with toys.PBDE congeners (predominantly BDE 47, 99, and 209) were detected in 100% of dust samples collected from 40 California daycare and preschool centers (Bradman et al. 2014). The mean and median totalPBDE (BDE 47, 99, 100, 118, 153, 154, 183, 190, 197, 203, 205, 206, 207, and 209) levels in dustsamples were 7,956.6 and 4,225 ng/g, respectively. Individual congeners (BDE 47, 99, 100 153, 154, and 209) were detected in indoor air samples at the facilities at mean levels ranging from 0.001 to 1.63 ng/m3. Hoffman et al. (2016) analyzed serum levels and handwipe samples from 83 children aged 12–36 monthsresiding in the state of North Carolina. Correlations between serum and handwipe levels of PBDEcongeners and behavioral patterns were observed. For example, increased age and increased activity was positively correlated to levels of PBDEs in serum and handwipe samples, while time spent sleeping (a measure of inactivity) was negatively correlated with PBDE levels in serum. It was reported that for eachadditional hour of sleep, BDE 47, BDE 99, BDE 100, and BDE 153 serum levels decreased by 12, 15, 9, and 10%, respectively, in the children. BDE 47, BDE 99, BDE 100, and BDE 153 levels in handwipesdecreased 30, 31, 30, and 23% for each additional hour of sleep. PBDEs 4396. POTENTIAL FOR HUMAN EXPOSURE6.7 POPULATIONS WITH POTENTIALLY HIGH EXPOSURES Workers who were involved in the production and manufacture of PBDE-containing plastics and plasticproducts were exposed to higher concentrations of PBDEs than the general population. Body burden dataindicate higher concentrations for workers exposed to PBDEs than for the general population. Occupational exposure to PBDEs also occurs in workers at plants that dismantle electronic equipment, computer monitor repair technicians, and automobile drivers, as well as other professions (Lindström1999). Occupational exposure occurs primarily by inhalation and ingestion of dust containing PBDEs. Stapleton et al. (2008) examined PBDE serum concentrations in workers involved with foam recyclingand carpet installation in the United States. Serum PBDE concentrations in foam recyclers (median,160 ng/g lipid) and carpet installers (median, 178 ng/g lipid) were significantly greater than a non-occupationally exposed control group (median, 19.3 ng/g lipid). Firefighters appear to have higher exposure potential to PBDEs and other types of flame retardantsbecause they are exposed to the combustion products of the flame retardants as well as the original forms of the chemicals. In a study conducted to examine exposure to PBDEs in 101 firefighters from Southern California, the median and geometric mean for total PBDEs (sum of BDE 28, 47, 99, 100, and 153) in theserum of the firefighters were 59.1 and 66.2 ng/g lipid, respectively (Park et al. 2015). These levels are approximately 40% greater when compared to the general population. The median and geometric mean for the same total PBDE congeners obtained from a subset of the NHANES survey with similar age andgender as the firefighters were 36 and 40.8 ng/g lipid, respectively. 6.8 ADEQUACY OF THE DATABASE Section 104(i)(5) of CERCLA, as amended, directs the Administrator of ATSDR (in consultation with theAdministrator of EPA and agencies and programs of the Public Health Service) to assess whetheradequate information on the health effects of PBDEs is available. Where adequate information is notavailable, ATSDR, in conjunction with NTP, is required to assure the initiation of a program of research designed to determine the health effects (and techniques for developing methods to determine such health effects) of PBDEs.The following categories of possible data needs have been identified by a joint team of scientists fromATSDR, NTP, and EPA. They are defined as substance-specific informational needs that if met would reduce the uncertainties of human health assessment. This definition should not be interpreted to mean PBDEs 4406. POTENTIAL FOR HUMAN EXPOSUREthat all data needs discussed in this section must be filled. In the future, the identified data needs will be evaluated and prioritized, and a substance-specific research agenda will be proposed.6.8.1 Identification of Data Needs Physical and Chemical Properties. Many of the relevant physical and chemical properties of the PBDEs are available (see Tables 4-3 and 4-4). Very limited data are available on the physical andchemical properties for the individual congeners (Braekevelt et al. 2003; Tittlemier et al. 2002).Important data, such as Kow, Koc, vapor pressure, and Henry’s Law constant, are necessary for theprediction of the environmental fate and transport of PBDEs.Production, Import/Export, Use, Release, and Disposal. According to the EmergencyPlanning and Community Right-to-Know Act of 1986, 42 U.S.C. Section 11023, industries are required to submit substance release and off-site transfer information to the EPA. The TRI, which contains thisinformation for 2014, became available in March 2016. This database is updated yearly and should provide a list of industrial production facilities and emissions.There are no current manufacturers of technical PBDEs in the United States. PentaBDE and octaBDEmixtures were voluntarily withdrawn from the U.S. marketplace by their manufacturers at the end of2004, leaving only decaBDE being marketed for use in commercial products in the United States (EPA2010). In December of 2009, the two remaining U.S. producers of decaBDE, Albemarle Corporation and Chemtura Corporation (formerly known as the Great Lakes Chemical Corporation), and the largest U.S. importer, ICL Industrial Products, Inc., announced commitments to phase out manufacture and importation of decaBDE for most uses in the United States by December 31, 2012, and to end manufacture and import for all uses by the end of 2013 (EPA 2013j). Many consumer goods enter theUnited States from other countries such as China. OctaBDE was never produced in China, and manufacture of the commercial pentaBDE mixture stopped in 2004; however, there are currently no restrictions on the use of decaBDE, which had a production volume of 20,500 metric tons in 2011 (Ni etal. 2013). It is unclear if items being treated with decaBDE are still entering U.S. markets from otherparts of the world.Given the importance assigned to dust ingestion as an exposure pathway to Americans, more data are needed on human bioavailability of PBDEs from external matrices, such as dust, as well as food for PBDEs 4416. POTENTIAL FOR HUMAN EXPOSUREexposure characterization. Characterization of the quantity of dust containing PBDEs, ingested byhumans, and in particular young children, would improve exposure estimates.Environmental Fate. Based on limited data, photolysis appears to be the dominant transformation process for some PBDEs (e.g., decaBDE) (Hua et al. 2003). PBDEs absorb light in the environmentalspectrum. Hua et al. (2003) found that decaBDE and the commercial octaBDE absorbed light up to 325 nm, which indicates that these compounds may be susceptible to photodegradation at environmentalwavelengths. However, the importance of photochemical transformation reactions in the environmentcannot be determined due to lack of quantitative rate information (EU 2002, 2003a). Better data on degradation via hydroxyl radical reaction and photolysis are needed. Based on a very limited number ofstudies, biodegradation does not appear to be significant for commercial mixtures of PBDEs (EU 2002, 2003a). Limited studies have been done on biodegradation of PBDEs in the environment under bothaerobic and anaerobic conditions, especially studies investigating dehalogenation mechanisms (EU 2002, 2003a). More studies are needed to determine conclusively if commercial PBDE mixtures, such as decaBDE, are degraded to lower-brominated congeners (e.g., BDE 47), which appear to bioaccumulate in fish, animals, and humans (see Section 6.4). Additional data on degradation via hydroxyl radical reaction and photolysis are needed. Since the toxicity and the environmental fate of PBDEs depend on specificPBDEs congeners, development of more data regarding congener-specific fate and transport of PBDEs inthe environment are needed.Bioavailability from Environmental Media. The absorption and distribution of PBDEs as a resultof inhalation, ingestion, and dermal exposure are discussed in Sections 3.4.1 and 3.4.2. PBDEs will existin both the vapor and particulate phase in both indoor and outdoor air, and more data are needed regarding the bioavailability of these substances in these two phases (Harrad et al. 2004) and the bioavailability of PBDEs from PBDE-contaminated toys (Chen et al. 2009).Food Chain Bioaccumulation. An abundance of monitoring data illustrates the uptake of lowerbrominated diphenyl ethers by aquatic organisms, which results in bioconcentration (see Section 6.4.4;Hardy 2002b). Congener components of the pentaBDE commercial product tend to bioconcentrate to different extents. DecaBDE and octaBDE commercial products do not bioconcentrate to the extent of thepenta mixtures; however, monitoring data clearly show that even the higher-brominated congeners are taken up. More information on bioaccumulation and biomagnification of PBDE and its congeners isneeded in assessing human health risks. PBDEs 4426. POTENTIAL FOR HUMAN EXPOSUREExposure Levels in Environmental Media. Reliable monitoring data for the levels of PBDEs incontaminated media at hazardous waste sites are needed so that the information obtained on levels ofPBDEs in the environment can be used in combination with the known body burden of PBDEs to assessthe potential risk of adverse health effects in populations living in the vicinity of hazardous waste sites.More monitoring data on the concentrations of total PBDEs and PBDE congeners in air in remote, rural, and urban areas, as well as areas near hazardous waste sites and incinerators, are needed. Althoughconcentrations are predicted to be low, monitoring data on PBDE concentrations in finished drinkingwater nationwide would be helpful. Sediment concentrations of PBDEs tend to be dominated by higherbrominated congeners (e.g., decaBDE or BDE 209) (deWit 2002; Dodder et al. 2002; Hale et al. 2001b, 2002). Monitoring data indicated that the concentrations of PBDEs are increasing in aquatic organisms with higher concentrations near point sources (Alaee et al. 1999; Dodder et al. 2000; Johnson and Olson 2001; Loganathan et al. 1995; Luross et al. 2000). Additional monitoring data on environmentalconcentrations of PBDEs would to useful to determine the extent of contamination in environmentalmedia, and also the mechanisms of human exposure to this class of chemicals.Exposure Levels in Humans. Body-burden data indicate that there are low-level exposures to lower-brominated PBDEs for the general population. Information about the average daily intake ofPBDEs is available (Bergman et al. 1999; EPA 2010; Lindström 1999; Lorber 2008; WHO 1994a). PBDE concentrations are reported in the current literature for serum, blood, breast milk, and adiposetissue of the general population and occupationally exposed individuals (CDC 2015; EPA 2010; WHO1994a). Additional data regarding the concentrations of PBDEs in body fluids or tissues of people who reside near hazardous waste sites are needed. This information is necessary for assessing the need toconduct health studies on these populations.Exposures of Children. The most important exposure pathway for infants to PBDEs likely occursthrough ingestion of breast milk (EPA 2010). PBDE intakes for infants and young children are typicallygreater than older children and adults (EPA 2010; Lorber 2008; Trudel et al. 2011; Wong et al. 2013). Since children tend to spend more time playing in carpeting, this leads to the potential to greater exposureto PBDEs through indoor dust. PBDEs have been detected in a variety of toys such as hard plastic toys(racing cars, vehicles, toy weapons, etc.), foam toys, rubber/soft plastic toys (dolls and teethers), and stuffed toys that were purchased in China from 2007 to 2008 (Chen et al. 2009). These findings indicateadditional possible exposure routes to children and toddlers through mouthing activities and dermalcontact with toys. Although there has been a gradual phase-out of pentaBDE, octaBDE, and now PBDEs 4436. POTENTIAL FOR HUMAN EXPOSUREdecaBDE, products are still in households that contain these substances. Therefore, continued biomonitoring data of infants and children is needed.Child health data needs relating to susceptibility are discussed in Section 3.12.2, Identification of DataNeeds: Children’s Susceptibility.Exposure Registries. No exposure registries for PBDEs were located. This substance is notcurrently one of the compounds for which a sub-registry has been established in the National Exposure Registry. The substance will be considered in the future when chemical selection is made for sub-registries to be established. The information that is amassed in the National Exposure Registry facilitatesthe epidemiological research needed to assess adverse health outcomes that may be related to exposure tothis substance.6.8.2 Ongoing Studies A prospective study of 316 mothers enrolled during pregnancy being conducted at the Icahn School ofMedicine at Mount Sinai, New York will examine prenatal exposure to complex mixtures of endocrine-disrupting compounds including PBDEs (RePORTER 2016).A study being conducted at the University of California, Berkeley is designed to examine exposures and health effects in vulnerable populations, such as pregnant women and children living in California, wherestricter flammability standards have resulted in very high flame retardant exposures (RePORTER 2016).The Center for Children's Environmental Health Research at the University of California, Berkeley willexamine novel methods of examining prenatal exposure to PBDEs and other compounds using sheddeciduous teeth and geographic information system (GIS) methods with remote sensing (RePORTER2016). A study at the University of Cincinnati is investigating two groups of persistent organic chemicals fortheir associations with adverse effects in child neurobehavior: PBDEs and perfluoroalkyl chemicals (including perfluorooctane sulfonic acid [PFOS] and perfluorooctanoic acid [PFOA]). The research project will provide novel information to the public about the developmental neurotoxicity of thesechemicals. It will also generate new data regarding PBDE exposure routes to aid in future prevention initiatives (RePORTER 2016). PBDEs 4446. POTENTIAL FOR HUMAN EXPOSUREA study is being conducted at the University of California, San Diego that is collecting and analyzingsediment and biota sample in the in the Southern California Bight for PBDEs (RePORTER 2016). PBDEs 4457. ANALYTICAL METHODSThe purpose of this chapter is to describe the analytical methods that are available for detecting,measuring, and/or monitoring PBDEs, their metabolites, and other biomarkers of exposure and effect to PBDEs. The intent is not to provide an exhaustive list of analytical methods. Rather, the intention is to identify well-established methods that are used as the standard methods of analysis. Many of theanalytical methods used for environmental samples are the methods approved by federal agencies andorganizations such as EPA and the National Institute for Occupational Safety and Health (NIOSH). Othermethods presented in this chapter are those that are approved by groups such as the Association ofOfficial Analytical Chemists (AOAC) and the American Public Health Association (APHA).Additionally, analytical methods are included that modify previously used methods to obtain lowerdetection limits and/or to improve accuracy and precision.PBDEs are analyzed in environmental and biological samples by methods quite similar to those used forPCBs (de Kok et al. 1977; Fries 1985b; Pomerantz et al. 1978). The analytical methods for PBDEs were developed relatively recently. There have been many advances in the technology and costs of analyticalinstruments used in the efforts directed at PBDE analysis. GC/MS with capillary columns (i.e., congenerspecific) is the primary analytical technique now used for PBDEs. Covaci et al. (2003) and Stapleton (2006) reviewed the determination of brominated flame retardants, with emphasis on PBDEs in environmental and human samples. The analysis methodology for PBDEsincludes several steps: sample collection and storage, sample pretreatment, extraction, cleanup and fractionation, and analytical determination. Care must be taken to assure that the sample collection follows quality-assurance protocols and that equipment and containers are free from contamination. It isimportant that laboratories utilize blanks when reporting trace concentrations of PBDEs. This practice will minimize the influence of trace contamination samples that can originate from a variety of sources.Most sample collections are by grab sampling; however, PBDEs may be concentrated from water ontosorbents. Desiccation of solid samples (e.g., soil, sediment, and sewage sludge) is largely done forconvenience. Dry samples are more efficiently homogenized, allowing for parallel determination of otheranalytes (e.g., lipid content) (Covaci et al. 2003).PBDEs are typically separated from the biological and environmental media by extraction with organic solvents. Liquid-solid extraction (e.g., Soxhlet apparatus) remains a widely used technique for solid PBDEs 4467. ANALYTICAL METHODSsamples despite recent advances in other extraction techniques. Typical solvents are hexane, toluene,hexane/acetone mixtures, or dichloromethane. New extractions techniques, such as accelerated solventextraction (ASE) or microwave-assisted extraction (MAE), are also currently used by a number oflaboratories. The advantage of these techniques is lower solvent consumption and reduced extraction time. Supercritical fluid extraction with solid-phase trapping has been used for the extraction ofbrominated flame retardants from sediment with CO2 as the supercritical fluid. Extraction with pressurized hot water (PHWE) has been used for the analysis of brominated analytes from sediment.Liquid-liquid extraction has been applied for river and seawater samples, using hexane/acetone mixtures.Solid-phase extraction has been used for the analysis of acidic and neutral brominated flame retardants from human plasma (Covaci et al. 2003). Cleanup steps are necessary to remove compounds that may interfere with the determination (e.g., humicacids, lipids) of PBDEs. Lipids (e.g., oils and fats) may be destroyed with concentrated sulfuric acid treatment either directly to the extract or using impregnated silica columns. Chromatography (e.g., gelpermeation, silica gel, Florisil) is used to remove other matrix interferences and to fractionate samples (Covaci et al. 2003).The identification and quantitation of PBDEs are most often accomplished by GC techniques. Capillaryor high-resolution gas chromatography (HRGC) columns capable of separating a substantial proportion ofthe congeners are indispensable, and GC detectors possessing high selectivity and sensitivity for the PBDEs are required. Historically, flame-ionization detectors (FID) or electron-capture detectors (ECD)were used. However, the MS detectors have become the main detection tool for PBDEs. MS detectorshave selectivity for PBDEs and can distinguish and individually measure homologs that may co-elute on a particular HRGC column. The use of MS is indispensable in the definitive identification of PBDEcongeners. One method of detection is ECNI as an ionization technique in combination with GC/MSanalysis (de Boer et al. 2000a). This method is advantageous because it offers a high sensitivity forcompounds with four or more bromine atoms. However, ECNI, although generally more sensitive and less costly than other ionization methods for PBDE analysis, does not provide information on themolecular ion cluster (as required for qualitative identification). It is also more subject to brominated interferences and does not allow the use of 13C-labeled standards for quantification (Ikonomou and Rayne2002). Conversely, electron ionization (EI) methods suffer from fragmentation of the molecular ions, creating difficulties in both identification and quantitation of congeners in full-scan and single ion monitoring (SIM) modes, respectively. For example, loss of bromine atoms from PBDE congenersduring EI may lead to incorrect identification of the parent ion as a lower-brominated congener. In PBDEs 4477. ANALYTICAL METHODSaddition, the relatively unpredictable fragmentation during EI restricts the utility of applying relativeresponse factors (RRFs) of one congener for which an analytical standard is available (e.g., 2,2,4,4’-tetraBDE or BDE 47) for other members of its homolog group (e.g., tetraBDEs). This can resultin either under- or overestimating concentrations of congeners for which analytical standards are notavailable (Ikonomou and Rayne 2002). In general, hepta- through decaBDE congeners are difficult todetermine accurately by GC analysis, especially in biological samples (Ikonomou and Rayne 2002). The analysis of BDE 209 and BDE 154 has some analytical difficulties. For example, BDE 209 (1) is notstable at high temperatures in the GC injector and GC column; (2) is sensitive to degradation by UV light (i.e., both sunlight and fluorescent light); (3) behaves differently in the MS source from those ofchlorinated and lower-brominated compounds (de Boer and Cofino 2002); and (4) may easily adsorb to small dust particles in the laboratory, which may result in sample contamination (Covaci et al. 2003). Thermal decomposition of BDE 209 can be avoided using a short GC column and a thermally inert GCinjection port (Beser et al. 2014). In contrast, BDE 154 usually co-elutes from most gas chromatographic columns with 2,2’,4,4’,5,5’-hexabromobiphenyl (PBB-153). In order to ensure the separation ofBDE 154 and PBB 153, analysts need to use a sufficiently long GC column. Thus, in order to accuratelydetermine the concentrations of BDE 209 and BDE 154 in analytical samples, analysts are requiredperform two separate GC measurements under different operating conditions. The difficulties in the determination of BDE 209 have been addressed through the use of liquid chromatography (LC)/MS/MS(Abdallah et al. 2009). The analytes, BDE 47, BDE 85, BDE 99, BDE 100, BDE 153, BDE 154, BDE 183, BDE 196, BDE 197, BDE 203, BDE 206, BDE 207, BDE 208, and BDE 209, were separatedand ionized in a single run using an isotope dilution method for indoor dust samples.7.1 BIOLOGICAL MATERIALSMethods for the determination of organobromine compounds such as PBDEs generally consist of thefollowing steps: extraction of the analyte from the sample matrix; cleanup to remove interferingcompounds; and analysis (separation and quantitation). The primary method of analysis is GC coupled with MS. Analytical methods have been developed for the determination of PBDEs in blood or serum, urine, feces, adipose tissue, liver, breast milk, and hair. The methods for determining PBDE residues inbiological samples are provided in Table 7-1. PBDEs 4487. ANALYTICAL METHODSTable 7-1. Analytical Methods for Determining Polybrominated Diphenyl Ethers (PBDEs) in Biological MaterialsSample matrix Preparation methodAnalytical methodSample detection limit Percent recovery ReferenceAnimal tissues (muscle, fat,and egg)Extraction with sulfuricacid; clean up with GPC/ silica column/carbon columnGC/MS (NCI)No data No data Sellström etal. 1993Animal serumExtraction with CH2Cl2 inhexane; H2SO4 to remove lipids; washed with NaOHfollowed by distilled water,then dehydrated through anhydrous Na2SO4ELISA (specific forBDE 27)0.2 µg/L 82–138 Ahn et al.2009Human adipose tissueSoxhlet extraction; clean up using 2 solid-phase extraction cartridgesCapillaryGC-EILRMS0.05–0.30 ng/g lipid81–103 Covaci et al.2002Human adipose tissueExtraction with methylene chloride; evaporate; cleanup on silica gel followed byclean up on alumina and on a carbon/silica gel columnHRGC/ HRMS0.73–120 pg/g No data Cramer et al.1990Human liver/ adipose tissueExtract with 2-propanol/ hexane; clean up with Lipidex 5000, column chromatography/GPCGC/MS (NCI)5 pg/g lipids 83 (54–116) liver;71 (51–95)adiposeMeironyté Guvenius etal. 2001Human milk Extract with potassiumoxalate/ethanol/diethylether/pentane; GPC; clean up on Florisil; elute with hexaneGC/MS (NCI/SIM)<0.6 ng/g fat No data WHO 1994aHuman milk Extract by column chromatography using hexane/dichloromethane/ hexane; clean up using GPCGC/MS (SIM)5 pg/g lipids 86–102 Meironyté etal. 1999Human milk Extract with n-hexane; HRGC- No data >80 Ohta et al.clean up using multi-layercolumnLRMS orLRGC2002HRMS (EISIM)Human plasmaExtract with formic acid,2-propanol, and water on a SPE column; derivatized GC/MS (NCI)1–10 pg/g plasma72 Thomsen etal. 2001using diazomethaneHuman serumExtraction with hexane/ MTBE (1:1); clean up silica gel/sulfuric acid columnGC-ECD;GC/MS (NCI)0.7 ng/g lipid weight69–104 (lowspike); 77–104 (high spike)Sjödin et al.1999a PBDEs 4497. ANALYTICAL METHODSTable 7-1. Analytical Methods for Determining Polybrominated Diphenyl Ethers (PBDEs) in Biological MaterialsSample matrix Preparation methodAnalytical methodSample detection limit Percent recovery ReferenceHuman serumExtraction with ethylacetate, evaporated todryness, and dissolved in DMSOELISA (specific forBDE 27)0.2 µg/L 83–90 Ahn et al.2009Human serumDenaturation with formicacid followed by SPEGC/MS 140–1,300 pg/g (lipid weight)40–106 Butt et al. 2016Human hair Physical extraction followed by washing with 0.3%polyoxyethylene laurylether and rinse with tap and distilled water;GC/MS 0.02 ng/g No data Malarvannan et al. 2013extraction using hexane/ dichloromethaneHuman hair Extraction with 4 N HCl andhexane (4:1); clean up on NaSO4/Florisil SPEcolumns (1:1); elute with hexaneGC/MS (SIM)0.025 pg/mg hair(lowerbrominated)2.5 pg/mg hair(BDE 209)70–90 Aleksa et al.2012aBDE = brominated diphenyl ether; DMSO = dimethyl sulfoxide; ECD = electron capture detection; EI = electron impact; EILR = electron impact low-resolution; ELISA = enzyme-linked immunosorbent assay; GC = gaschromatography; GPC = gel permeation chromatography; HRGC = high resolution gas chromatography;HRMS = high resolution mass spectrometry; LRGC = low resolution gas chromatography; LRMS = low resolution mass spectrometry; MS = mass spectrometry; MTBE = methyl-tert-butyl either; NCI = negative chemical ionization;SIM = selected ion monitoring; SPE = solid phase extraction PBDEs 4507. ANALYTICAL METHODSDetermination of hydroxylated PBDEs metabolites were reported by Malmberg et al. (2005). Themethylated derivatives of PBDE-OH metabolites were identified by GC-MS ECNI and EI analysis afterextraction from the sample matrix, cleanup to remove interfering compounds, and methylation of thehydroxyl group. In plasma samples, the internal standard, BDE 138, had 71 and 98% recovery (relativestandard deviation, 8 and 2%) on days 1 and 5 after dosing, respectively. Residues in biological samples can be extracted using sulfuric acid, 2-propanol/hexane, methylenechloride, n-hexane, formic acid/2-propanol/water, or hexane/methyl t-butyl ether (Cramer et al. 1990;Meironyté Guvenius et al. 2001; Ohta et al. 2002; Sellström et al. 1993; Sjödin et al. 1999a; Thomsen etal. 2001). Samples are cleaned up to remove interferences using Florisil, silica gel, alumina or activated-charcoal column chromatography, gel permeation chromatography (GPC), and/or LC (Cramer et al. 1990;Meironyté Guvenius et al. 2001; Sellström et al. 1993; Sjödin et al. 1999a). Most techniques are based on analysis by GC coupled with MS (WHO 1994a). Capillary columns and temperature programming allowthe separation of the different PBDE congeners. High recoveries (69–104%) of PBDE residues are obtained by the available analytical methods. Typically, the limit of quantitation for PBDE residues isabout 0.7 ng/g lipid in blood serum, 5 pg/g lipid in human milk, 0.3 ng/g lipid in adipose tissue, and 0.025 pg/mg in hair (Aleksa et al. 2012a; Covaci et al. 2002; Meironyté Guvenius et al. 1999; Sjödin et al. 1999a). Additionally, a selective competitive enzyme-linked immunosorbent assay (ELISA) has been developed to detect BDE 47 (Ahn et al. 2009). This method also reports high recoveries (83–90%) and alimit of quantitation for BDE 47 in blood of 0.2 µg/L.7.2 ENVIRONMENTAL SAMPLES Most environmental analyses have been performed using multi-residue methods involving solventextraction of the analytes from the sample matrix, cleanup to remove interfering compounds, determination by GC with confirmation using MS. New methods and technologies are evolving, and thishas resulted in lower detection limits. Analytical methods for the determination of PBDEs in environmental samples are given in Table 7-2. Like PCBs, air samples containing PBDEs are usually collected by pumping air through a samplercontaining a glass-fiber filter and adsorbent trap to separate the particle-bound and vapor-phase fractions,respectively (Dodder et al. 2000; Hillery et al. 1997). The filters and adsorbents are then Soxhletextracted with acetone/hexane, and the extracts are cleaned up and analyzed by high-resolution GCtechniques. Beser et al. (2014) discussed a GC/MS method to quantify PBDEs that used microwave PBDEs 4517. ANALYTICAL METHODSTable 7-2. Analytical Methods for Determining Polybrominated Diphenyl Ethers(PBDEs) in Environmental SamplesSample Analytical Sample PercentMatrix Preparation method method detection limit recovery ReferenceAir Air pumped through glassfiber filter and adsorbenttrap; filters and adsorbentsare Soxhlet extracted with acetone/hexane; cleaned-up by column chromatographyAir Samples collected using a large-volume active samplerat 30 m3/hour for 24 hoursonto filters followed bymicrowave assisted extraction using hexane/acetone.Air Passive samplers set outwith PUF disks; PUF disks were collected at 10-dayintervals over 50 days; active samplers had low volume pump with PUF plugshoused in a glass holder;solvent extraction (collected hexane layer); washed withH2SO4 back extraction using dimethyl sulfoxide followed by column elution containing Florisil, hexane, and anhydrous Na2SO4Dust Samples collected fromvacuum cleaner bags;extracted using hexane byan accelerated solventextraction system,concentrated, treated with H2SO4, liquid/liquid backextraction using dimethylsulfoxide, column elution Dust Samples collected fromvacuum cleaner bags;extracted using microwave-assisted solvent extraction(collected hexane layer);washed with H2SO4, then deionized water, and dried with anhydrous Na2SO4GC/MS No data No data Dodder et al.2000GC/MS 0.063– 80–120% Beser et al.0.210 pg/m3 2014GC/MS No data 41–78% Hazrati and (passive Harrad 2007samplers)42–80%(activesamplers)GC/MS 0.03 ng/g No data Harrad et al.2006ELISA 0.2 µg/L 105±22.7 Ahn et al. 2009(specific forBDE 47) PBDEs 4527. ANALYTICAL METHODSTable 7-2. Analytical Methods for Determining Polybrominated Diphenyl Ethers(PBDEs) in Environmental SamplesSample Analytical Sample PercentMatrix Preparation method method detection limit recovery ReferenceWater Clean up by GFF followed byPUF plugs; extraction with dichloromethaneWater Clean up by disk-type C18solid-phase extractionSewage Soxhlet extractionacetone:hexane (1:1, v/v),clean-up with liquid-liquidextraction with fumingsulfuric acid, GPC and silica gel column, and a basicalumina columnSewage Extract with chloroform;evaporate and dissolve residue in ethanolSediment Extract with hexane anddichloromethane (1:1 v/v),treat with copper, sulfonatewith sulfuric acid, clean andfraction using neutral, acid,and alkaline silica gelchromatographySediment Clean up by cartridge-typeFlorosil extractionSediment Pressurized hot waterextraction coupled with clean up by LCSediment Extract with acetone; clean up on FlorisilFish Extract with dichloromethane:n-hexane (1:1, v/v) column chromatography on silicaand Al2O3Fish Extract with acetone-hexane + hexane-ethyl ether;treatment with sulfuric acidor clean up on alumina;chromatography on silica gelGC/MSCapillaryGC-ECDHRGC/MSGC/MSGC/MSCapillaryGC-ECDLC-GC/MS/ FIDNAA;GC/ECGC/MSGC/EC; GC/MS0.2–1.4 pg/L No data Yang et al. 20140.12 pg/L 103±8.6 (river Yamamoto et al.water); 199787±10.7 (sea water)No data No data Zennegg et al.20130.06 μg/g No data WHO 1994a 1.93–227 pg/g 88.8−138 Tang et al. 20149.7 ng/g 91±6.3 Yamamoto et al.19970.71 ng/g No data Kuosmanen etal. 2002<5 ng/g; <5 ng/g No data Watanabe et al.1987No data 58–106 Yang et al. 20080.1 μg/g fat No data Andersson and Blomkvist 1981 PBDEs 4537. ANALYTICAL METHODSTable 7-2. Analytical Methods for Determining Polybrominated Diphenyl Ethers(PBDEs) in Environmental SamplesSample Matrix Preparation methodAnalytical methodSample detection limitPercentrecovery ReferenceFish SE, ASE, and MAEperformed followed byevaporation to dryness;dissolved in n-hexane;HRGC/ HRMS24.8 pg/g 79–118 (SE)50–96 (MAE)Wang et al.2010added silica gel, column filtered; concentrated; silicagel column followed by a basic alumina column eluted with n-hexane/ dichloromethane (1:1, v/v)Fish Extract with dichloromethane GC-HRMSon chromatography column; (NCI)clean-up using GPC;fractionation using silica gelcolumn5–93 pg/g No data Alaee et al.2001Fish Extract clean up with GPCand mini-column chromatography; concentrationGC/MS (NCI)0.01–0.2 ng/g lipid88–128 Akutsu et al.2001Animal tissuesHomogenize; extract with GC/MS n-hexane-acetone; treatment (NCI)with sulfuric acid; GPC;chromatography or silica gelchromatography or activated charcoal10 pg/g No data Jansson et al.1991Vegetables Homogenize; MAE; extract GC/MSclean up with Florisil or silicacartridge; elute withn-hexane:toluene (80:20,v/v)1-3 ng/g 99–106 Bizkarguenaga et al. 2014Vegetables Extract with acetone/ n-hexane (1:1 v/v)GC/MS (ion 1 ng/kg drytrap) weight82–98 (mean) Parolini et al.2012ASE = accelerated solvent extraction; BDE = brominated diphenyl ether; EC = electron capture; ECD = electroncapture detection; ELISA = enzyme-linked immunosorbent assay; GC = gas chromatography; GFF = glass fibrefilter; GPC = gel permeation chromatography; HRGC = high resolution gas chromatography; HRMS = high resolution mass spectrometry; LC = liquid chromatography; MAE = microwave-assisted extraction; MS = mass spectrometry;NAA = neutron activation analysis; NCI = negative chemical ionization; PUF = polyurethane foam; SE = soxhletextraction PBDEs 4547. ANALYTICAL METHODSassisted extraction (MAE) rather than the traditional Soxhlet extraction technique in order to shorten theextraction time and quantity of solvents used. For a sampling rate of 30 m3 per hour over a samplingduration of 24 hours, they achieved low limits of quantification (0.063 pg/m3for BDE 28, BDE 49, BDE 47, BDE 66, and BDE 100; 0.105 pg/m3 for BDE 119 and BDE 99; and 0.210 pg/m3 for BDE 155, BDE 154, BDE 153, BDE 139, and BDE 183). Sampling of PBDEs is also performed using passive ordiffusive samplers (Covaci et al. 2003). Hazrati and Harrad (2007) describe passive sampling with polyurethane foam disks (PUF) of BDE 28, BDE 47, BDE 99, and BDE 100 with mean recoveries of57 and 62% for passive and active samplers, respectively. Harrad and Hunter (2006) performed passiveair sampling with PUF disks in the United Kingdom. BDE 28, BDE 47, BDE 99, BDE 100, BDE 153, and BDE 154 were detected using GC/MS. The detection limit for individual BDEs were approximately0.05 pg/m3.Passive air sampling techniques have been developed to monitor both vapor- and particulate-phase PBDEs in indoor air through the use of a PUF disk and glass fiber filter (GFF) sampling media (Abdallah 2010). The PUF disks and GFFs were evaluated independently by initial soxhlet extraction with dichloromethane, concentration of extracts, and purification with SPE cartridge. Elution with hexane:dichloromethane (1:1, v/v), evaporation, reconstitution in methanol, and analysis usingLC/MS/MS followed. BDE 47, BDE 85, BDE 99, BDE 100, BDE 153, BDE 154, BDE 183, BDE 196, BDE 197, BDE 203, BDE 206, BDE 207, BDE 208, and BDE 209 were evaluated using this technique. The octaBDE to decaBDE analytes were detected on the GFF media only, indicating that these congeners are expected to primarily be found in the particulate phase.Residues in environmental samples can be extracted using chloroform, acetone, acetone-hexane, hexaneacetone, and hexane-ether (Andersson and Blomkvist 1981; Jansson et al. 1991; Watanabe et al. 1987;WHO 1994a). Samples are cleaned up to remove interferences using Florisil, silica gel, alumina oractivated charcoal column chromatography, GPC, and/or LC (Akutsu et al. 2001; Alaee et al. 2001b;Andersson and Blomkvist 1981; Jansson et al. 1991; Watanabe et al. 1987; Yamamoto et al. 1997). Vegetable and soil samples have been prepared for analysis using, focused ultrasound solid-liquidextraction (Bizkarguenaga et al. 2014).As for biological samples, quantitation of environmental samples is also usually done by GC. Capillarycolumns are required for the separation of the individual congeners in a mixture (WHO 1994a). High recoveries (88–128%) of PBDE residues in environmental samples are obtained by the availableanalytical methods (Akutsu et al. 2001). Typically, the limit of quantitation for PBDE residues is about PBDEs 4557. ANALYTICAL METHODS0.12 ng/mL in water, 9.7 μg/kg in sediment, and 0.2 μg/kg lipid in fish (Akutsu et al. 2001; Yamamoto etal. 1997). The first inter-laboratory study on PBDEs in environmental samples showed that there is good agreement for quantification of BDE 47 and BDE 100 congeners. Additionally, a selective competitiveELISA has been developed to detect BDE 47 in dust (Ahn et al. 2009). This method also reports high recoveries (105%) and a limit of quantitation for BDE 47 in blood of 0.2 µg/L.7.3 ADEQUACY OF THE DATABASE Section 104(i)(5) of CERCLA, as amended, directs the Administrator of ATSDR (in consultation with theAdministrator of EPA and agencies and programs of the Public Health Service) to assess whetheradequate information on the health effects of PBDEs is available. Where adequate information is notavailable, ATSDR, in conjunction with NTP, is required to assure the initiation of a program of research designed to determine the health effects (and techniques for developing methods to determine such health effects) of PBDEs.The following categories of possible data needs have been identified by a joint team of scientists fromATSDR, NTP, and EPA. They are defined as substance-specific informational needs that if met would reduce the uncertainties of human health assessment. This definition should not be interpreted to meanthat all data needs discussed in this section must be filled. In the future, the identified data needs will be evaluated and prioritized, and a substance-specific research agenda will be proposed. 7.3.1 Identification of Data Needs Methods for Determining Biomarkers of Exposure and Effect.Exposure. Methods used as biomarkers for exposure to PBDEs are available (Ahn et al. 2009; Aleksa etal. 2012a; Brilliant et al. 1978; Covaci et al. 2002; Eyster et al. 1983; Landrigan et al. 1979; MeironytéGuvenius et al. 1999; Sjödin et al. 1999a; Wolff et al. 1982). Analytical methods of sufficient precision and accuracy are presently available for the determination of PBDEs in adipose tissue, serum, breast milk,and hair (Ahn et al. 2009; Aleksa et al. 2012a; Burse et al. 1980; Covaci et al. 2002; Domino et al. 1980;Fawkes et al. 1982; Fehringer 1975a; Meironyté Guvenius et al. 1999; Sjödin et al. 1999a; Willet et al.1978; Wolff et al. 1979a, 1979b). Additional congener standards are needed for PBDEs analysis. Only30–40 congener standards are currently available for identification and quantification of PBDEs (Eljarratet al. 2002; Sjödin et al. 1998). Metabolites are also important biomarkers for exposure to PBDEs. Ryden et al. (2012) discussed a GC/MS method for the analysis of hydroxylated PBDE metabolites in PBDEs 4567. ANALYTICAL METHODShuman blood. GC/MS has been used to identify hydroxylated-BDE metabolites from recombinantcytochrome P450 by Simpson et al. (2015), and Gross et al. (2015) used GC/MS and GC/MS/MS.Effect. No studies have been conducted to determine if known effects of PBDEs exposure can bequantitatively correlated with PBDE exposure. Methods for Determining Parent Compounds and Degradation Products in EnvironmentalMedia. Analytical methods of sufficient sensitivity are presently available for the determination ofPBDEs in environmental samples (Akutsu et al. 2001; Andersson and Blomkvist 1981; Covaci et al. 2003; Stapleton 2006; Yamamoto et al. 1997). Methods for determining degradation products and metabolites of PBDE are needed. There is noinformation in the literature of detectable biodegradation of PBDEs in the environment under aerobic oranaerobic conditions. The analysis of PBDE pyrolysis degradation products, such as PBDD/PBDF, isoften disturbed by the presence of PBDEs. Ebert et al. (1999) demonstrated that by using a Florisilcolumn in a sample clean-up process, almost complete separation of PBDEs and PBDDs/PBDFs isachieved before analysis by GC/MS.7.3.2 Ongoing Studies Analysis of PBDEs and other anthropogenic pollutants in marine mammals by a novel application ofcomprehensive two-dimensional gas chromatography with time- of-flight mass spectrometry(GCxGC/TOF-MS) is being studied at the University of California, San Diego (RePORTER 2016). PBDEs 4578. REGULATIONS, ADVISORIES, AND GUIDELINESMRLs are substance specific estimates, which are intended to serve as screening levels, are used byATSDR health assessors and other responders to identify contaminants and potential health effects thatmay be of concern at hazardous waste sites.ATSDR has derived an intermediate-duration inhalation MRL of 0.006 mg/m3 for lower-brominatedPBDEs based on a NOAEL for changes in thyroid hormones in rats (Great Lakes Chemical Corporation 2000).ATSDR has derived an acute-duration oral MRL of 0.00006 mg/kg/day for lower-brominated PBDEsbased on a LOAEL for endocrine effects in rat dams and reproductive and neurobehavioral effects in F1 offspring from a series of reports (Kuriyama et al. 2005, 2007; Talsness et al. 2005). ATSDR has derived an intermediate-duration oral MRL of 0.000003 mg/kg/day for lower-brominated PBDEs based on a minimal LOAEL for decreased testosterone in male rats (Zhang et al. 2013b).ATSDR has derived an acute-duration oral MRL of 0.01 mg/kg/day for decaBDE based on a NOAEL forneurobehavioral effects in mice (Johansson et al. 2008). ATSDR has derived an intermediate-duration oral MRL of 0.0002 mg/kg/day for decaBDE based on a minimal LOAEL for increased serum glucose in rats (Zhang et al. 2013a).IARC has classified PBDE as a Group 3 carcinogen (not classifiable as to its carcinogenicity to humans) based on inadequate evidence of carcinogenicity in humans and inadequate or limited evidence inexperimental animals (IARC 2014). The EPA assigns the cancer category Group D (not classifiable as to human carcinogenicity) to mono-, di-, tri-, tetra-, penta-, hexa-, octa-, and nonaBDEs (IRIS 2003a, 2003b, 2003c, 2003d, 2003e, 2004, 2005, 2006) and reports “inadequate information” to classify the specific congeners 2,2’,4,4’-tetraBDE, 2,2’,4,4’,5-pentaBDE, and 2,2’,4,4’,5,5’-hexaBDE (IRIS 2008b, 2008c,2008d). However, EPA assigns a classification of “suggestive evidence of carcinogenic potential” for decaBDE (IRIS 2008a). The Department of Health and Human Services has not classified PBDEs as carcinogens (NTP 2011). ACGIH has no data regarding cancer classifications for PBDEs (ACGIH 2014).The EPA’s reference doses (RfDs) for penta-, octa-, and decaBDEs are 2x10-3, 3x10-3, and 7x10-3 mg/kg/day, respectively (IRIS 2003c, 2004, 2008a). For the specific congeners2,2’,4,4’-tetraBDE, 2,2’,4,4’, 5-pentaBDE, and 2,2’,4,4’,5,5’-hexaBDE, the RfDs are 1x10-4, 1x10-4, and PBDEs 4588. REGULATIONS, ADVISORIES, AND GUIDELINES2x10-4 mg/kg/day, respectively (IRIS 2008b, 2008c, 2008d). No reference concentrations (RfCs) were derived for PBDEs.OSHA has not set PELs to protect workers against adverse health effects resulting from exposure toPBDEs (OSHA 2013a, 2013b). No guidelines for worker exposure limits have been recommended byACGIH (2014) or NIOSH (2014).WHO has not established any air quality guidelines for PBDEs (WHO 2010). PBDEs are not designatedas hazardous air pollutants, and no acute exposure guidelines (AEGLs) have been derived (EPA 2013a, 2014a). The Department of Energy (DOE) has established values for responding to potential releases ofairborne monoBDE for use in community emergency planning. The values established by the DOE(2012) are the Protective Active Criteria (PAC-1, -2, and -3). The PAC-1, -2, and -3 values are 0.29, 3.2, and 19 mg/m3, respectively, and represent increasing severity of effects (mild, irreversible, and life threatening, respectively) for a 1-hour exposure (DOE 2012). The American Industrial HygieneAssociation (AIHA) has no Emergency Response Planning Guidelines (ERPGs-1, -2, -3) for PBDEs(AIHA 2014).WHO has not established any drinking water guidelines for PBDEs (WHO 2011) and the EPA has not setdrinking water standards for PBDEs (EPA 2009a, 2009b, 2012, 2013c, 2013d, 2014c). The FDA has notset allowable levels for PBDEs in bottled water (FDA 2013).Under the Toxic Substances Control Act (TSCA), mono-, penta-, octa-, and decaBDEs are on the list ofchemicals that manufacturers and importers must report for each plant site at which they manufactured orimported PBDEs during the reporting period specified (EPA 1998a). MonoBDE (represented by CASRegistry Number 101-55-3) has been designated as a hazardous substance pursuant to CERCLA of 1980 (EPA 2013g). The owner and operator of any facility that produces, uses, or stores a CERCLA hazardoussubstance is required to immediately report releases to any environmental media, if the amount released is equal to or exceeds the specified “reportable quantity” assigned to the substance. The reportable quantityfor monoBDE is 100 pounds (45 kg) (EPA 2013g). However, PBDEs are no longer manufactured orimported in the United States as of January 2014 (EPA 2013j)The Emergency Planning and Community Right-to-Know Act (EPCRA) has identified decaBDE as atoxic chemical and monoBDE as hazardous waste, and the Master Testing list includes penta-, octa-, and decaBDEs (EPA 2006, 2013e, 2014d). MonoBDE is on the Resource Conservation and Recovery Act PBDEs 4598. REGULATIONS, ADVISORIES, AND GUIDELINES(RCRA) waste minimization persistent, bioaccumulative, and toxic (PBT) priority chemical list and the groundwater monitoring list (EPA 1998b, 2013f). The international and national regulations, advisories, and guidelines regarding PBDEs in air, water, and other media are summarized in Table 8-1. PBDEs 4608. REGULATIONS, ADVISORIES, AND GUIDELINESTable 8-1. Regulations, Advisories, and Guidelines Applicable to PBDEsAgency Description Information ReferenceINTERNATIONALGuidelines:IARC Carcinogenicity classification IARC 2014DecaBDE Group 3a WHO Air quality guidelines No data WHO 2010Drinking water quality guidelines No data WHO 2011NATIONALRegulations and Guidelines:a. AirACGIH TLV-TWA No data ACGIH 2014AIHA ERPGs No data AIHA 2014DOE PACs DOE 2012MonoBDEPAC-1 0.29 mg/m3PAC-2 3.2 mg/m3PAC-3 19 mg/m3 EPA AEGLs No data EPA 2013aRegulated toxic and flammable No data EPA 2013bsubstances under Section 112(r) of the 40 CFR 68.130Clean Air ActHazardous Air Pollutants No data EPA 2014aNAAQS No data EPA 2014bNIOSH REL No data NIOSH 2014IDLH No dataOSHA PEL (8-hour TWA) for general industry No data OSHA 2013a29 CFR 1910.1000,Table Z-1Highly hazardous chemicals No data OSHA 2013b29 CFR 1910.119,Appendix Ab. WaterEPA Designated as hazardous substances No data EPA 2013cin accordance with Section 311(b)(2)(A) 40 CFR 116.4of the Clean Water ActDrinking water contaminant candidate No data EPA 2009alist 74 FR 51850Drinking water standards and health No data EPA 2012advisoriesNational primary drinking water No data EPA 2009bregulations PBDEs 4618. REGULATIONS, ADVISORIES, AND GUIDELINESTable 8-1. Regulations, Advisories, and Guidelines Applicable to PBDEsAgency Description Information ReferenceNATIONAL (cont.)National recommended water qualitycriteriaNo data EPA 2014cReportable quantities of hazardoussubstances designated pursuant to Section 311 of the Clean Water ActNo data EPA 2013d40 CFR 117.3c. FoodFDA Allowable levels for contaminants in No data FDA 2013bottled water 21 CFR 165.110EAFUSb No data FDA 2014d. OtherACGIH Carcinogenicity classification No data ACGIH 2014EPA Carcinogenicity classificationMonoBDEDiBDEsTriBDEsTetraBDEsPentaBDEsHexaBDEsOctaBDEsNonaBDEsDecaBDE2,2’,4,4’-tetraBDE2,2’,4,4’,5-pentaBDE2,2’,4,4’,5,5’-hexaBDEGroup Dc Group Dc Group Dc Group Dc Group Dc Group Dc Group Dc Group Dc Suggestive evidence of carcinogenicpotentialInadequate informationInadequate informationInadequate informationIRIS 2006IRIS 2005IRIS 2003eIRIS 2003dIRIS 2004IRIS 2003aIRIS 2003cIRIS 2003bIRIS 2008aIRIS 2008bIRIS 2008cIRIS 2008dRfC No dataRfDPentaBDEsOctaBDEsDecaBDE2,2’,4,4’-tetraBDE2,2’,4,4’,5-pentaBDE2,2’,4,4’,5,5’-hexaBDE2x10-3 mg/kg/day3x10-3 mg/kg/day7x10-3 mg/kg/day1x10-4 mg/kg/day1x10-4 mg/kg/day2x10-4 mg/kg/dayIRIS 2004IRIS 2003aIRIS 2008aIRIS 2008bIRIS 2008cIRIS 2008dChemical substances subject to proposed or final TSCA rules or ordersMonoBDEPentaBDEsTSCA Section 5(a)(2)TSCA Section 4EPA 1998aOctaBDEs TSCA Section 4DecaBDE TSCA Section 4EPCRA Section 313 Toxic Chemicals EPA 2006DecaBDE Yes PBDEs 4628. REGULATIONS, ADVISORIES, AND GUIDELINESTable 8-1. Regulations, Advisories, and Guidelines Applicable to PBDEsAgency Description Information ReferenceNATIONAL (cont.)Identification and listing of hazardouswasteMonoBDE YesEPA 2013e40 CFR 261,Appendix VIIIMaster Testing ListPentaBDEs YesEPA 2014dOctaBDEs YesDecaBDE YesRCRA waste minimization PBT prioritychemical listEPA 1998b63 FR 60332MonoBDE YesStandards for owners and operators ofhazardous waste TSD facilities;groundwater monitoring listMonoBDE YesEPA 2013f40 CFR 264,Appendix IXSuperfund, emergency planning, and community right-to-knowDesignated CERCLA hazardoussubstance and reportable quantityMonoBDEEPA 2013g40 CFR 302.4Statutory codeFinal RQ pounds2,4100Effective date of toxic chemical EPA 2013hrelease reportingDecaBDE 1/1/8740 CFR 372.65Extremely hazardous substancesand its threshold planning quantityNo data EPA 2013i40 CFR 355,Appendix ADHHS Carcinogenicity classification No data NTP 2011aGroup 3: Not classifiable as to its carcinogenicity to humansbThe EAFUS list of substances contains ingredients added directly to food that FDA has either approved as food additives or listed or affirmed as GRAS.cGroup D: Not classifiable as to human carcinogenicityACGIH = American Conference of Governmental Industrial Hygienists; AEGL = acute exposure guideline levels;AIHA = American Industrial Hygiene Association; BDE = brominated diphenyl ether; CERCLA = Comprehensive Environmental Response, Compensation, and Liability Act; CFR = Code of Federal Regulations;DHHS = Department of Health and Human Services; DOE = Department of Energy; EAFUS = Everything Added to Food in the United States; EPA = Environmental Protection Agency; EPCRA = Emergency Planning and CommunityRight-To-Know Act; ERPG = emergency response planning guidelines; FDA = Food and Drug Administration;FR = Federal Register; GRAS = generally recognized as safe; IARC = International Agency for Research on Cancer;IDLH = immediately dangerous to life or health; IRIS = Integrated Risk Information System; NAAQS = NationalAmbient Air Quality Standards; NIOSH = National Institute for Occupational Safety and Health; NTP = NationalToxicology Program; OSHA = Occupational Safety and Health Administration; PAC = protective action criteria;PBDE = polybrominated diphenyl ether; PBT = persistent, bioaccumulative, and toxic; PEL = permissible exposure limit; RCRA = Resource Conservation and Recovery Act; REL = recommended exposure limit; RfC = inhalation reference concentration; RfD = oral reference dose; RQ = reportable quantity; TLV = threshold limit values;TSCA = Toxic Substances Control Act; TSD = treatment, storage, and disposal; TWA = time-weighted average;WHO = World Health Organization PBDEs 4639. 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GLOSSARYAbsorption—The taking up of liquids by solids, or of gases by solids or liquids.Acute Exposure—Exposure to a chemical for a duration of 14 days or less, as specified in theToxicological Profiles.Adsorption—The adhesion in an extremely thin layer of molecules (as of gases, solutes, or liquids) to the surfaces of solid bodies or liquids with which they are in contact.Adsorption Coefficient (Koc)—The ratio of the amount of a chemical adsorbed per unit weight oforganic carbon in the soil or sediment to the concentration of the chemical in solution at equilibrium.Adsorption Ratio (Kd)—The amount of a chemical adsorbed by sediment or soil (i.e., the solid phase)divided by the amount of chemical in the solution phase, which is in equilibrium with the solid phase, at afixed solid/solution ratio. It is generally expressed in micrograms of chemical sorbed per gram of soil orsediment.Benchmark Dose (BMD)—Usually defined as the lower confidence limit on the dose that produces aspecified magnitude of changes in a specified adverse response. For example, a BMD10 would be thedose at the 95% lower confidence limit on a 10% response, and the benchmark response (BMR) would be10%. The BMD is determined by modeling the dose response curve in the region of the dose responserelationship where biologically observable data are feasible.Benchmark Dose Model—A statistical dose-response model applied to either experimental toxicological or epidemiological data to calculate a BMD.Bioconcentration Factor (BCF)—The quotient of the concentration of a chemical in aquatic organismsat a specific time or during a discrete time period of exposure divided by the concentration in thesurrounding water at the same time or during the same period.Biomarkers—Broadly defined as indicators signaling events in biologic systems or samples. They have been classified as markers of exposure, markers of effect, and markers of susceptibility.Cancer Effect Level (CEL)—The lowest dose of chemical in a study, or group of studies, that producessignificant increases in the incidence of cancer (or tumors) between the exposed population and itsappropriate control.Carcinogen—A chemical capable of inducing cancer.Case-Control Study—A type of epidemiological study that examines the relationship between a particular outcome (disease or condition) and a variety of potential causative agents (such as toxic chemicals). In a case-controlled study, a group of people with a specified and well-defined outcome isidentified and compared to a similar group of people without outcome.Case Report—Describes a single individual with a particular disease or exposure. These may suggestsome potential topics for scientific research, but are not actual research studies.Case Series—Describes the experience of a small number of individuals with the same disease orexposure. These may suggest potential topics for scientific research, but are not actual research studies. PBDEs 52810. GLOSSARYCeiling Value—A concentration of a substance that should not be exceeded, even instantaneously.Chronic Exposure—Exposure to a chemical for 365 days or more, as specified in the ToxicologicalProfiles.Cohort Study—A type of epidemiological study of a specific group or groups of people who have had acommon insult (e.g., exposure to an agent suspected of causing disease or a common disease) and arefollowed forward from exposure to outcome. At least one exposed group is compared to one unexposed group.Cross-sectional Study—A type of epidemiological study of a group or groups of people that examinesthe relationship between exposure and outcome to a chemical or to chemicals at one point in time.Data Needs—Substance-specific informational needs that if met would reduce the uncertainties of human health assessment.Developmental Toxicity—The occurrence of adverse effects on the developing organism that may resultfrom exposure to a chemical prior to conception (either parent), during prenatal development, orpostnatally to the time of sexual maturation. Adverse developmental effects may be detected at any pointin the life span of the organism.Dose-Response Relationship—The quantitative relationship between the amount of exposure to atoxicant and the incidence of the adverse effects.Embryotoxicity and Fetotoxicity—Any toxic effect on the conceptus as a result of prenatal exposure toa chemical; the distinguishing feature between the two terms is the stage of development during which theinsult occurs. The terms, as used here, include malformations and variations, altered growth, and in uterodeath.Environmental Protection Agency (EPA) Health Advisory—An estimate of acceptable drinking waterlevels for a chemical substance based on health effects information. A health advisory is not a legallyenforceable federal standard, but serves as technical guidance to assist federal, state, and local officials.Epidemiology—Refers to the investigation of factors that determine the frequency and distribution ofdisease or other health-related conditions within a defined human population during a specified period. Genotoxicity—A specific adverse effect on the genome of living cells that, upon the duplication ofaffected cells, can be expressed as a mutagenic, clastogenic, or carcinogenic event because of specific alteration of the molecular structure of the genome.Half-life—A measure of rate for the time required to eliminate one half of a quantity of a chemical fromthe body or environmental media.Immediately Dangerous to Life or Health (IDLH)—The maximum environmental concentration of a contaminant from which one could escape within 30 minutes without any escape-impairing symptoms orirreversible health effects.Immunologic Toxicity—The occurrence of adverse effects on the immune system that may result fromexposure to environmental agents such as chemicals. PBDEs 52910. GLOSSARYImmunological Effects—Functional changes in the immune response.Incidence—The ratio of individuals in a population who develop a specified condition to the totalnumber of individuals in that population who could have developed that condition in a specified timeperiod. Intermediate Exposure—Exposure to a chemical for a duration of 15–364 days, as specified in the Toxicological Profiles.In Vitro—Isolated from the living organism and artificially maintained, as in a test tube.In Vivo—Occurring within the living organism.Lethal Concentration(LO) (LCLO)—The lowest concentration of a chemical in air that has been reportedto have caused death in humans or animals.Lethal Concentration(50) (LC50)—A calculated concentration of a chemical in air to which exposure fora specific length of time is expected to cause death in 50% of a defined experimental animal population.Lethal Dose(LO) (LDLo)—The lowest dose of a chemical introduced by a route other than inhalation thathas been reported to have caused death in humans or animals.Lethal Dose(50) (LD50)—The dose of a chemical that has been calculated to cause death in 50% of adefined experimental animal population.Lethal Time(50) (LT50)—A calculated period of time within which a specific concentration of a chemicalis expected to cause death in 50% of a defined experimental animal population.Lowest-Observed-Adverse-Effect Level (LOAEL)—The lowest exposure level of chemical in a study,or group of studies, that produces statistically or biologically significant increases in frequency or severityof adverse effects between the exposed population and its appropriate control.Lymphoreticular Effects—Represent morphological effects involving lymphatic tissues such as the lymph nodes, spleen, and thymus.Malformations—Permanent structural changes that may adversely affect survival, development, orfunction.Minimal Risk Level (MRL)—An estimate of daily human exposure to a hazardous substance that is likely to be without an appreciable risk of adverse noncancer health effects over a specified route and duration of exposure.Modifying Factor (MF)—A value (greater than zero) that is applied to the derivation of a Minimal RiskLevel (MRL) to reflect additional concerns about the database that are not covered by the uncertaintyfactors. The default value for a MF is 1.Morbidity—State of being diseased; morbidity rate is the incidence or prevalence of disease in a specific population.Mortality—Death; mortality rate is a measure of the number of deaths in a population during a specified interval of time. PBDEs 53010. GLOSSARYMutagen—A substance that causes mutations. A mutation is a change in the DNA sequence of a cell’s DNA. Mutations can lead to birth defects, miscarriages, or cancer.Necropsy—The gross examination of the organs and tissues of a dead body to determine the cause ofdeath or pathological conditions.Neurotoxicity—The occurrence of adverse effects on the nervous system following exposure to a chemical.No-Observed-Adverse-Effect Level (NOAEL)—The dose of a chemical at which there were nostatistically or biologically significant increases in frequency or severity of adverse effects seen betweenthe exposed population and its appropriate control. Effects may be produced at this dose, but they are notconsidered to be adverse.Octanol-Water Partition Coefficient (Kow)—The equilibrium ratio of the concentrations of a chemicalin n-octanol and water, in dilute solution.Odds Ratio (OR)—A means of measuring the association between an exposure (such as toxic substances and a disease or condition) that represents the best estimate of relative risk (risk as a ratio of the incidenceamong subjects exposed to a particular risk factor divided by the incidence among subjects who were notexposed to the risk factor). An OR of greater than 1 is considered to indicate greater risk of disease in the exposed group compared to the unexposed group.Organophosphate or Organophosphorus Compound—A phosphorus-containing organic compound and especially a pesticide that acts by inhibiting cholinesterase.Permissible Exposure Limit (PEL)—An Occupational Safety and Health Administration (OSHA)allowable exposure level in workplace air averaged over an 8-hour shift of a 40-hour workweek.Pesticide—General classification of chemicals specifically developed and produced for use in the controlof agricultural and public health pests.Pharmacokinetics—The dynamic behavior of a material in the body, used to predict the fate(disposition) of an exogenous substance in an organism. Utilizing computational techniques, it providesthe means of studying the absorption, distribution, metabolism, and excretion of chemicals by the body.Pharmacokinetic Model—A set of equations that can be used to describe the time course of a parentchemical or metabolite in an animal system. There are two types of pharmacokinetic models: data-basedand physiologically-based. A data-based model divides the animal system into a series of compartments,which, in general, do not represent real, identifiable anatomic regions of the body, whereas thephysiologically-based model compartments represent real anatomic regions of the body.Physiologically Based Pharmacodynamic (PBPD) Model—A type of physiologically based dose-response model that quantitatively describes the relationship between target tissue dose and toxic endpoints. These models advance the importance of physiologically based models in that they clearlydescribe the biological effect (response) produced by the system following exposure to an exogenoussubstance. Physiologically Based Pharmacokinetic (PBPK) Model—Comprised of a series of compartmentsrepresenting organs or tissue groups with realistic weights and blood flows. These models require a PBDEs 53110. GLOSSARYvariety of physiological information: tissue volumes, blood flow rates to tissues, cardiac output, alveolarventilation rates, and possibly membrane permeabilities. The models also utilize biochemicalinformation, such as air/blood partition coefficients, and metabolic parameters. PBPK models are alsocalled biologically based tissue dosimetry models.Prevalence—The number of cases of a disease or condition in a population at one point in time. Prospective Study—A type of cohort study in which the pertinent observations are made on eventsoccurring after the start of the study. A group is followed over time.q1*—The upper-bound estimate of the low-dose slope of the dose-response curve as determined by themultistage procedure. The q1* can be used to calculate an estimate of carcinogenic potency, the incremental excess cancer risk per unit of exposure (usually μg/L for water, mg/kg/day for food, and μg/m3 for air).Recommended Exposure Limit (REL)—A National Institute for Occupational Safety and Health(NIOSH) time-weighted average (TWA) concentration for up to a 10-hour workday during a 40-hourworkweek.Reference Concentration (RfC)—An estimate (with uncertainty spanning perhaps an order ofmagnitude) of a continuous inhalation exposure to the human population (including sensitive subgroups)that is likely to be without an appreciable risk of deleterious noncancer health effects during a lifetime.The inhalation reference concentration is for continuous inhalation exposures and is appropriatelyexpressed in units of mg/m3 or ppm.Reference Dose (RfD)—An estimate (with uncertainty spanning perhaps an order of magnitude) of thedaily exposure of the human population to a potential hazard that is likely to be without risk of deleteriouseffects during a lifetime. The RfD is operationally derived from the no-observed-adverse-effect level(NOAEL, from animal and human studies) by a consistent application of uncertainty factors that reflectvarious types of data used to estimate RfDs and an additional modifying factor, which is based on aprofessional judgment of the entire database on the chemical. The RfDs are not applicable tononthreshold effects such as cancer.Reportable Quantity (RQ)—The quantity of a hazardous substance that is considered reportable underthe Comprehensive Environmental Response, Compensation, and Liability Act (CERCLA). Reportablequantities are (1) 1 pound or greater or (2) for selected substances, an amount established by regulation either under CERCLA or under Section 311 of the Clean Water Act. Quantities are measured over a 24-hour period.Reproductive Toxicity—The occurrence of adverse effects on the reproductive system that may resultfrom exposure to a chemical. The toxicity may be directed to the reproductive organs and/or the relatedendocrine system. The manifestation of such toxicity may be noted as alterations in sexual behavior, fertility, pregnancy outcomes, or modifications in other functions that are dependent on the integrity ofthis system.Retrospective Study—A type of cohort study based on a group of persons known to have been exposed at some time in the past. Data are collected from routinely recorded events, up to the time the study is undertaken. Retrospective studies are limited to causal factors that can be ascertained from existingrecords and/or examining survivors of the cohort.Risk—The possibility or chance that some adverse effect will result from a given exposure to a chemical. PBDEs 53210. GLOSSARYRisk Factor—An aspect of personal behavior or lifestyle, an environmental exposure, or an inborn orinherited characteristic that is associated with an increased occurrence of disease or other health-relatedevent or condition.Risk Ratio—The ratio of the risk among persons with specific risk factors compared to the risk amongpersons without risk factors. A risk ratio greater than 1 indicates greater risk of disease in the exposedgroup compared to the unexposed group.Short-Term Exposure Limit (STEL)—The American Conference of Governmental IndustrialHygienists (ACGIH) maximum concentration to which workers can be exposed for up to 15 minutescontinually. No more than four excursions are allowed per day, and there must be at least 60 minutesbetween exposure periods. The daily Threshold Limit Value-Time Weighted Average (TLV-TWA) maynot be exceeded.Standardized Mortality Ratio (SMR)—A ratio of the observed number of deaths and the expected number of deaths in a specific standard population.Target Organ Toxicity—This term covers a broad range of adverse effects on target organs orphysiological systems (e.g., renal, cardiovascular) extending from those arising through a single limited exposure to those assumed over a lifetime of exposure to a chemical.Teratogen—A chemical that causes structural defects that affect the development of an organism.Threshold Limit Value (TLV)—An American Conference of Governmental Industrial Hygienists (ACGIH) concentration of a substance to which most workers can be exposed without adverse effect. The TLV may be expressed as a Time Weighted Average (TWA), as a Short-Term Exposure Limit(STEL), or as a ceiling limit (CL).Time-Weighted Average (TWA)—An allowable exposure concentration averaged over a normal 8-hourworkday or 40-hour workweek.Toxic Dose(50) (TD50)—A calculated dose of a chemical, introduced by a route other than inhalation, which is expected to cause a specific toxic effect in 50% of a defined experimental animal population.Toxicokinetic—The absorption, distribution, and elimination of toxic compounds in the living organism.Uncertainty Factor (UF)—A factor used in operationally deriving the Minimal Risk Level (MRL) orReference Dose (RfD) or Reference Concentration (RfC) from experimental data. UFs are intended to account for (1) the variation in sensitivity among the members of the human population, (2) theuncertainty in extrapolating animal data to the case of human, (3) the uncertainty in extrapolating fromdata obtained in a study that is of less than lifetime exposure, and (4) the uncertainty in using lowest-observed-adverse-effect level (LOAEL) data rather than no-observed-adverse-effect level (NOAEL) data.A default for each individual UF is 10; if complete certainty in data exists, a value of 1 can be used;however, a reduced UF of 3 may be used on a case-by-case basis, 3 being the approximate logarithmicaverage of 10 and 1.Xenobiotic—Any chemical that is foreign to the biological system. PBDEs A-1APPENDIX A. ATSDR MINIMAL RISK LEVELS AND WORKSHEETSThe Comprehensive Environmental Response, Compensation, and Liability Act (CERCLA) [42 U.S.C. 9601 et seq.], as amended by the Superfund Amendments and Reauthorization Act (SARA) [Pub. L. 99– 499], requires that the Agency for Toxic Substances and Disease Registry (ATSDR) develop jointly withthe U.S. Environmental Protection Agency (EPA), in order of priority, a list of hazardous substances mostcommonly found at facilities on the CERCLA National Priorities List (NPL); prepare toxicological profiles for each substance included on the priority list of hazardous substances; and assure the initiationof a research program to fill identified data needs associated with the substances.The toxicological profiles include an examination, summary, and interpretation of available toxicologicalinformation and epidemiologic evaluations of a hazardous substance. During the development oftoxicological profiles, Minimal Risk Levels (MRLs) are derived when reliable and sufficient data exist toidentify the target organ(s) of effect or the most sensitive health effect(s) for a specific duration for agiven route of exposure. An MRL is an estimate of the daily human exposure to a hazardous substancethat is likely to be without appreciable risk of adverse noncancer health effects over a specified durationof exposure. MRLs are based on noncancer health effects only and are not based on a consideration ofcancer effects. These substance-specific estimates, which are intended to serve as screening levels, are used by ATSDR health assessors to identify contaminants and potential health effects that may be ofconcern at hazardous waste sites. It is important to note that MRLs are not intended to define clean-up oraction levels.MRLs are derived for hazardous substances using the no-observed-adverse-effect level/uncertainty factorapproach. They are below levels that might cause adverse health effects in the people most sensitive tosuch chemical-induced effects. MRLs are derived for acute (1–14 days), intermediate (15–364 days), and chronic (365 days and longer) durations and for the oral and inhalation routes of exposure. Currently, MRLs for the dermal route of exposure are not derived because ATSDR has not yet identified a methodsuitable for this route of exposure. MRLs are generally based on the most sensitive chemical-induced end point considered to be of relevance to humans. Serious health effects (such as irreparable damage to the liver or kidneys, or birth defects) are not used as a basis for establishing MRLs. Exposure to a levelabove the MRL does not mean that adverse health effects will occur.MRLs are intended only to serve as a screening tool to help public health professionals decide where tolook more closely. They may also be viewed as a mechanism to identify those hazardous waste sites that PBDEs A-2APPENDIX Aare not expected to cause adverse health effects. Most MRLs contain a degree of uncertainty because ofthe lack of precise toxicological information on the people who might be most sensitive (e.g., infants,elderly, nutritionally or immunologically compromised) to the effects of hazardous substances. ATSDRuses a conservative (i.e., protective) approach to address this uncertainty consistent with the public healthprinciple of prevention. Although human data are preferred, MRLs often must be based on animal studiesbecause relevant human studies are lacking. In the absence of evidence to the contrary, ATSDR assumes that humans are more sensitive to the effects of hazardous substance than animals and that certain persons may be particularly sensitive. Thus, the resulting MRL may be as much as 100-fold below levels that have been shown to be nontoxic in laboratory animals.Proposed MRLs undergo a rigorous review process: Health Effects/MRL Workgroup reviews within theDivision of Toxicology and Human Health Sciences, expert panel peer reviews, and agency-wide MRLWorkgroup reviews, with participation from other federal agencies and comments from the public. Theyare subject to change as new information becomes available concomitant with updating the toxicologicalprofiles. Thus, MRLs in the most recent toxicological profiles supersede previously published levels.For additional information regarding MRLs, please contact the Division of Toxicology and Human Health Sciences, Agency for Toxic Substances and Disease Registry, 1600 Clifton Road NE, Mailstop F-57, Atlanta, Georgia 30329-4027. PBDEs A-3APPENDIX AMINIMAL RISK LEVEL (MRL) WORKSHEETChemical Name: Polybrominated Diphenyl Ethers (PBDEs)[lower-brominated diphenyl ethers]CAS Numbers: 32536-52-0 (octaBDE)Date: March 2017Profile Status: FinalRoute: [X] Inhalation [ ] OralDuration: [ ] Acute [X] Intermediate [ ] ChronicGraph Key: 2Species: RatMinimal Risk Level: 0.006 [ ] mg/kg/day [ ] ppm [X] mg/m3 Reference: Great Lakes Chemical Corporation. 2000. A 90-day inhalation toxicity study ofoctabromodiphenyl oxide in albino rats, dated 04/04/02. Submitted to the U.S. Environmental ProtectionAgency under TSCA Section 8E. OTS0574171-1.Experimental design: This is an unpublished study in which a commercial octaBDE product (Lot No. 9525DA23B, bromine content 78.7%, composition and purity not otherwise specified) was administered to groups of 10 male and 10 female Crl:CD(SD)IGS BR rats, via nose-only inhalation as a dust aerosol, in measured concentrations of 0 (filtered air-only), 1.1, 16, or 202 mg/m3 for 6 hours/day, 5 days/week, for13 weeks. The mean MMADs in the low to high level groups were 2.0, 2.7, and 2.8 microns; thecorresponding mean GSDs were 3.37, 3.72, and 3.01. Clinical and physical signs, body weight, foodconsumption, and survival were evaluated throughout the study. Ophthalmic, hematology (11 indices), serum chemistry (18 indices), and serum thyroid hormone (TSH, total T3, and total T4) evaluations were performed near the end of the exposure period. Urinalyses were not conducted. Comprehensivenecropies, organ weight measurements, and histological examinations (including respiratory tract and thyroids) were performed following exposure termination. Effects noted in study and corresponding doses: Hepatic, nasal, lung, thyroid, and ovarian effects wereobserved. The liver was affected in both sexes as shown by dose-related increases in centrilobularhepatocellular hypertrophy at ≥16 mg/m3 and liver weight (absolute and relative) at 202 mg/m3. Totalincidences of centrilobular hepatocellular hypertrophy in the 0, 1.1, 16, and 202 mg/m3 groups were1/10 (minimal), 0/10, 3/10 (all minimal), and 10/10 (6 minimal, 2 mild, 2 moderate) in males, and 0/10, 0/10, 3/10 (all minimal), and 6/10 (3 minimal, 3 mild) in females. Changes in nasal goblet cells wereincreased at 202 mg/m3, but showed no clear dose-related increasing trends for incidence or severity.Total incidences of goblet cell hypertrophy (minimal or mild) were slightly increased in nasal level II ofboth sexes at ≥1.1 mg/m3; incidences in 0, 1.1, 16, and 202 mg/m3 exposure groups were 4/10 (allminimal), 9/10 (7 minimal, 2 mild), 6/10 (all minimal), and 10/10 (9 minimal, 1 mild) in males, and2/10 (all minimal), 6/10 (all minimal), 4/10 (all minimal), and 8/10 (all minimal) in females. Goblet cell hypertrophy was also slightly increased in nasal level IV in males at 202 mg/m3 (4/10, 0/10, 1/10, and 8/10, all minimal severity, not increased in females). Histological changes in the lungs included alveolarhistiocytosis and chronic active inflammation that were only clearly induced at 202 mg/m3. Totalincidences of alveolar histiocytosis at 0, 1.1, 16, and 202 mg/m3 were 3/10 (2 mild, 1 minimal), 5/10 (allminimal), 5/10 (all minimal), and 10/10 (5 minimal, 3 mild, 2 moderate) in males, and 0/10, 5/10 (allminimal), 2/10 (all minimal), and 10/10 (1 minimal, 7 mild, 2 moderate) in females. Corresponding totalincidences of chronic active lung inflammation were 0/10, 0/10, 2/10 (both minimal), and 10/10 (5 minimal, 4 mild, 1 moderate) in males, and 0/10, 1/10 (minimal), 1/10 (minimal), and 10/10 (2 minimal, 5 mild, 3 moderate) in females. Gross lung changes also occurred in both sexes at202 mg/m3; these included lung firmness and white discoloration and/or enlargement in the bronchial PBDEs A-4APPENDIX Aand/or mediastinal lymph nodes. The lymph node effects correlated with the histological finding ofgranulomatous inflammation. There were no exposure-related gross or histopathological changes in thespleen, bone marrow, thymus, or other tissues, including thyroid. Thyroid hormone assessments, however, showed exposure-related decreases in mean thyroxine (total T4) at ≥16 mg/m3 in both sexes and increases in TSH at ≥16 mg/m3 in males and 202 mg/m3 in females. The changes were usuallystatistically significant (p<0.05 or p<0.01) compared to controls and were considered to be consistentwith chemical-induced hypothyroidism. There were no serum T3 changes. Qualitative histological evaluations of step sections of ovaries showed an absence of corpora lutea in 3/10 females at 202 mg/m3, compared to 0/10 in the control and lower exposure groups. This 30% incidence was interpreted to be atreatment-related effect because an absence of corpora lutea was considered unusual in rats at 20 weeks ofage.Other findings included some hematological alterations in 202 mg/m3 females that were not considered tobe exposure-related (slightly increased mean activated partial thromboplastin time, and decreased mean corpuscular hemoglobin and mean corpuscular hemoglobin concentration without effects on red bloodcell counts, hematocrit, or hemoglobin levels). Serum chemistry evaluations showed that cholesterol wassignificantly increased (66.2% more than controls, p<0.01) in 202 mg/m3 females, but the magnitude ofthe elevation was not considered toxicologically significant. Some other statistically significant serumchemistry alterations (increased mean globulin and total protein, decreased albumin/globulin ratio) also occurred in the 202 mg/m3 females, but were not considered exposure-related due to small magnitudes ofchanges and lack of similar findings in the males.Dose and end point used for MRL derivation: 1.1 mg/m3[X] NOAEL [ ] LOAELConsidering the unclear adversity of minimal severity goblet cell hypertrophy, lack of clear dose-relatedincreasing trends for incidence and severity of this nasal effect, identification of both a NOAEL(1.1 mg/m3) and LOAEL (16 mg/m3) for changes in thyroid hormones, and abundant evidence for thyroid effects of PBDEs in oral studies, the NOAEL for effects on thyroid hormones is the most appropriatebasis for derivation of the MRL.Uncertainty factors used in MRL derivation:[X] 3 for extrapolation from animals to humans with dosimetric adjustments[X] 10 for human variability[X] 3 modifying factor for incomplete databaseWas a conversion factor used from ppm in food or water to a mg/body weight dose? Not applicable.Was a conversion used from intermittent to continuous exposure? The NOAEL was adjusted tocontinuous exposure as follows: 1.1 mg/m3 x 6 hours/24 hours x 5 days/7 days = 0.196 mg/m3 If an inhalation study in animals, list conversion factors used in determining human equivalent dose: Thehuman equivalent NOAEL (NOAELHEC) was calculated from the duration-adjusted NOAEL (NOAELADJ) using EPA RfC methodology as follows:NOAELHEC = NOAELADJ x RDDR = 0.196 mg/m3 x 2.7 = 0.53 mg/m3The RDDR for the extrathoracic region was used to extrapolate deposited doses in rats to deposited dosesin humans. The following parameters were used to calculate the RDDR of 2.7: MMAD of 2.0 μm with a PBDEs A-5APPENDIX Amean GSD (sigma g) of 3.37; default human body weight of 70 kg, and a default female F344 rat body weight of 180 g.Based on these values, the MRL for lower brominated diphenyl ethers is derived as follows:MRL = NOAELHEC ÷ (UF x MF) = 0.53 ÷ (30 x 3) = 0.006 mg/m3 Other additional studies or pertinent information that lend support to this MRL: This is the onlyintermediate-duration inhalation study of PBDEs.The thyroid is a sensitive target of lower-brominated BDEs in orally exposed animals. A LOAEL forreduced serum T4 hormone levels in rat dams that were exposed to 2,2’,4,4’,5-pentaBDE (BDE 99)(Kuriyama et al. 2007) was used as a co-critical end point for the basis for the acute oral MRL for lowerbrominated BDEs. This study is supported by numerous studies that report reduced serum T4 levels inadult, nonpregnant mice and rats following acute exposure to commercial pentaBDE mixtures (Bromkal 70, Bromkal 70-5 DE, DE-71), and the commercial octaBDE mixture DE-79, or2,2’,4,4’-tetraBDE (BDE 47), indicating significant reductions of 19–92% following gavage exposure atdoses ≥10 and ≥0.8 mg/kg/day in rats and mice, respectively, for 1–14 days (Darnerud and Sinjari 1996;Fowles et al. 1994; Hallgren and Darnerud 1998, 2002; Hallgren et al. 2001;Hoppe and Carey 2007;Richardson et al. 2008; Stoker et al. 2004, 2005; Zhou et al. 2001). In developing animals, numerousstudies have reported decreased serum T4 and/or T3 levels in pups after gestational and lactationalexposure to commercial pentaBDE mixtures (DE-71, Bromkal 70-5 DE), BDE 99, or BDE 47 at doses as low as 0.3 mg/kg/day in rats and 452 mg/kg/day in mice (Blanco et al. 2013; Bondy et al. 2011, 2013;Ellis-Hutchings et al. 2006; Kodavanti et al. 2010; Kuriyama et al. 2007; Miller et al. 2012; Poon et al. 2011; Shah et al. 2011; Skarman et al. 2005; Szabo et al. 2009; Wang et al. 2011a; Zhou et al. 2002).Hepatic effects observed in critical study also support the selected point of departure (POD), as a dose-related increased centrilobular hepatocellular hypertrophy was observed in males and females exposed tooctaBDE at concentrations ≥16 mg/m3; however, this end point was not selected as a co-critical effect, as the increase in incidence was only significant at 202 mg/m3 (Great Lakes Chemical Corporation 2000).Hydroxylated metabolites (OH-PBDEs) may be responsible for the toxic action of PBDEs. There is some evidence that the CYPs involved in metabolism of PBDEs, as well as the OH-PBDEs formed, are different in rats and humans (Erratico et al. 2011, 2012, 2013)—see Section 3.4.3 for more information. However, there is no available evidence indicating that the complement of OH-PBDEs formed in humansis any more or less potent than the complement of OH-PBDEs formed in rats. Therefore, the current use of an animal-to-human uncertainty factor is appropriate for calculating an MRL based on effects observedin rats.Agency Contact (Chemical Manager): Hana Pohl PBDEs A-6APPENDIX AMINIMAL RISK LEVEL (MRL) WORKSHEETChemical Name: Polybrominated Diphenyl Ethers (PBDEs)[lower-brominated diphenyl ethers]CAS Numbers: 60348-60-9 (2,2’,4,4’,5-pentaBDE)Date: March 2017Profile Status: FinalRoute: [ ] Inhalation [X] OralDuration: [X] Acute [ ] Intermediate [ ] ChronicGraph Key: 24, 60, 63Species: RatMinimal Risk Level: 0.00006 (6x10-5) [X] mg/kg/day [ ] ppm [ ] mg/m3 References: Kuriyama SN, Talsness CE, Grote K, et al. 2005. Developmental exposure to low dosePBDE 99: effects on male fertility and neurobehavior in rat offspring. Environ Health Perspect113(2):149-154. Talsness CE, Shakibaei M, Kuriyama SN, et al. 2005. Ultrastructural changes observed in rat ovaries following in utero and lactational exposure to low doses of a polybrominated flame retardant. ToxicolLett 157(3):189-202.Kuriyama SN, Wanner A, Fidalgo-Neto AA, et al. 2007. Developmental exposure to low-dose PBDE99: Tissue distribution and thyroid hormone levels. Toxicology 242(1-3):80-90.Experimental design: In the first study (Kuriyama et al. 2005), pregnant rats (16–20/group) were given asingle dose of 2,2’,4,4’,5-pentaBDE (BDE 99, 98% purity) at 0, 0.06, or 0.3 mg/kg via gavage in peanutoil vehicle on GD 6. Dams were allowed to deliver, and litter size was not artificially altered. Emergence of postnatal reflexes and developmental landmarks (eruption of incisors, fur development, eye opening, and testes descent) was evaluated in all pups (163–200/group). Locomotor activity was evaluated over24-hour periods on PNDs 36 and 71 in one male and female per litter (16–20 litters/group). F1 maleoffspring were sacrificed as adults (~PND 140, 12 males/group) and the thymus, spleen, liver, testis,epididymis, seminal vesicle, and ventral prostate were weighed. The right testis and caudal epididymis were retained for spermatid and sperm counts and morphology, respectively. Additionally, blood wascollected for analysis of testosterone and LH levels. In 15–19 F1 males/group, reproductive function wasassessed at ~PND 150. F1 males were mated with untreated females in a 1:1 ratio for 14 days. Theability of males to impregnate unexposed females was assessed, and pregnant dams were sacrificed onGD 21 for assessment of the number of implantations, resorptions, and fetuses in the F2 generation. Uterine and fetal weight was recorded, and fetuses were sexed and examined for external anomalies. In a separate group of F1 males, male sexual behavior was assessed in 20 males/group at ~PND 160. F1males were mated with untreated females in estrus (1:1) and the sexual behavior of each mating was recorded for 20 minutes.In the second study (Talsness et al. 2005), pregnant rats (14–17/group) were exposed to BDE 99 according to the exposure protocol for Study 1. The F1 offspring were weaned on PND 22. The femaleoffspring were necropsied in estrus (based on vaginal cytology) on approximately PND 90. Histological evaluation of the ovary (10/group), uterus (5–7/group), and vagina (5–9/group) was performed. Ovarian follicles were counted in 10 ovaries from each group. One ovary from one female offspring in each group was analyzed by transmission electron microscopy. Twenty virgin F1 females per group were mated with non-exposed males to evaluate fertility. The F1 dams were sacrificed on GD 21 and the uterus was excised. The uterine and F2 fetal weights and the number of implantations, resorptions, and fetuses were PBDEs A-7APPENDIX Adetermined. The F2 fetuses were examined for external anomalies and when present, they were stained and examined for skeletal anomalies.In the third study (Kuriyama et al. 2007), pregnant rats (15–20/group) were exposed to BDE 99 according to the exposure protocol for Study 1. On PND 1, approximately half of the dams (8–10/group) and theiroffspring were sacrificed. Liver samples were collected for enzyme activity (EROD, UDPGT) and bloodwas collected for determination of thyroid hormones (T3, free-T3, T4, free-T4). In pups, blood and livertissue were pooled by gender on a litter basis. On PND 14, 2 pups/sex/litter (7–11 litters/group) weresacrificed, and liver samples and blood were collected for analysis. On PND 22, remaining dams (7– 11/group) and 2 pups/sex/litter were sacrificed, and liver samples and blood were collected for analysis.Effects noted in study and corresponding doses: Study 1 (Kuriyama et al. 2005): No exposure-related effects were observed for the age at fur developmentor eye opening, testes descent, or the ability to master the rotating rod test. However, significant delays inthe eruption of incisors in F1 pups and the development of the cliff-drop aversion reflex were observed in F1 males in the 0.3 mg/kg group, compared with controls. Total activity, time spent active, the duration of activity per active phase, and the total activity per active phase were all significantly increased in F1offspring on PND 36 in the 0.3 mg/kg group, compared with controls. On PND 71, the increased totalactivity and time spent active persisted in the 0.3 mg/kg group, and was also significantly increased in the0.06 mg/kg group. In the group sacrificed on PND 140, no exposure-related changes were observed inbody weight, liver weight, or thymus weight; however, absolute spleen weight was significantly increased by 9% in the 0.06 and 0.3 mg/kg groups, and relative spleen weight was significantly increased by 12% in the 0.06 mg/kg group. Compared with controls, significantly altered male reproductive organ weights atPND 140 included a 10 and 11% decrease in relative testes and epididymis weight, respectively, in the0.3 mg/kg group and a 5% decrease in relative epididymis weight in the 0.06 mg/kg group; no significantchanges were observed in absolute organ weights. In both dose groups, the number of spermatids and sperm and daily sperm production were significantly decreased, compared with controls. No exposure-related effects were observed for sperm morphology. No changes were observed in serum testosterone orLH levels. Despite sperm alterations, no significant exposure-related effects were observed in male reproductive function or the majority of male sexual behaviors. The only significantly altered male sexual behavior was a 32% decrease in the percent of males with two or more ejaculations.Study 2 (Talsness et al. 2005): No statistically significant, exposure-related histological changes were observed at the light microscopic level in the ovary, uterus, or vagina of female offspring, and no exposure-related effects were observed in the number of ovarian follicles. However, multipleultrastructural changes were noted in the ovaries of PND 90 female offspring from dams exposed to0.06 or 0.3 mg/kg, including destruction of the surface of the serosal epithelial cells, necrosis, andnumerous vesicular structures with dense granular material within the cytoplasm. Additional changes observed in the 0.3 mg/kg group included degenerative changes and aggregates of small and largevesicles filled with homogeneously dense granular material in the cytoplasm and clumped chromatinwithin the condensed nucleus. No exposure-related changes were found for F1 female pregnancy rate,total implantation sites, implantation sites/dam, F2 fetuses/gravid dam, or total number of live F2 fetuses. However, the resorption rates were 12 and 15% in the 0.06 and 0.3 mg/kg groups, respectively, compared with the control rate of 9%. Statistics were not reported; however, the resorption rates in the exposed rats were also reportedly increased compared with historical controls (average control resorption rate=5.4%,with rates up to 10% considered to be within normal limits). In addition, the percentage of litters with resorptions was higher in the exposed females, being 47% in the control group and 69 and 72% in the0.06 and 0.3 mg/kg groups, respectively. In F2 pups, mean fetal weight was significantly increased by5% in the 0.06 mg/kg group, but not in the 0.3 mg/kg group, compared with controls. Three fetuses fromdifferent litters in the 0.3 mg/kg/day group showed skeletal anomalies (tail, skull, vertebrae); however, PBDEs A-8APPENDIX Athis incidence of anomalies in 3/18 litters is not significantly elevated compared with the control incidence of 0/19 (Fisher’s exact test, performed for this review).Study 3 (Kuriyama et al. 2007): Serum T4 levels were significantly decreased by 23–33% in the 0.06 and 0.3 mg/kg dams, sacrificed on PND 1. No changes were observed in T3, free-T3, or free-T4 at PND 1 orany thyroid hormone levels at PND 22 in dams. In pups, no dose-related changes were observed atPND 1 or 14. At PND 22, serum T4 was significantly decreased by in F1 males and females and serumfree-T4 was significantly decreased in F1 females from the 0.3 mg/kg group (19–23% reductions). Hepatic EROD activity was significantly decreased in PND 22 dams from the 0.3 mg/kg group; no otherchanges in hepatic enzyme activity were observed in dams. In F1 offspring, hepatic UDPGT activity was significantly increased in females at PND 1 and EROD activity was significantly elevated in males atPND 22; no other changes in hepatic enzyme activity were observed in F1 offspring.Dose and end point used for MRL derivation: 0.06 mg/kg[ ] NOAEL [X] LOAELCollectively, these studies indicate a LOAEL of 0.06 mg/kg for endocrine effects in F0 dams (reduced serum T4) and reproductive and neurobehavioral effects in F1 adult offspring (impaired spermatogenesis, ultrastructural changes in ovaries, increased resorptions in F1 females mated to unexposed males, andincreased spontaneous motor activity). A NOAEL was not identified.Uncertainty factors used in MRL derivation:[X] 10 for extrapolation from a LOAEL to a NOAEL[X] 10 for extrapolation from animals to humans[X] 10 for human variabilityWas a conversion factor used from ppm in food or water to a mg/body weight dose? Not applicable(gavage studies).Was a conversion used from intermittent to continuous exposure? Not applicable (single exposure studies).If an inhalation study in animals, list conversion factors used in determining human equivalent dose: Not applicable.Other additional studies or pertinent information that lend support to this MRL: Support for reproductive effects in F1 animals as a co-critical end point: In a companion study to thecritical studies described above, pregnant rats (8/group) were administered 2,2´,4,4´-tetrabromodiphenylether (BDE 47, 98% purity) at 0, 0.14, or 0.7 mg/kg via gavage in peanut oil vehicle on GD 6 (Talsness etal. 2008). As observed in pentaBDE-exposed F1 females, ultrastructural changes (accumulation ofvesicular structures with homogeneously dense granular material in the cytoplasm of the stromal cells,large vacuoles) were observed in the ovaries of F1 females from both dose groups on PND 100. No exposure-related changes were observed in F1 female fertility or F2 litter parameters. F1 males were notevaluated for developmental reproductive effects following tetraBDE exposure.Support for altered open-field activity in F1 animals as a co-critical end point: Alterations in open-fieldactivity have been consistently reported in mice exposed to BDE 99 at doses ≥0.8 mg/kg on PND 3 or 10 and evaluated at 2–8 months of age, characterized by decreased activity during the first 20-minute period PBDEs A-9APPENDIX Afollowed by increased activity during the third 20-minute period (Eriksson et al. 2002b, 2006; Fischer etal. 2008; Sand et al. 2004; Viberg et al. 2002, 2004a, 2004b). These findings indicate an initial decrease in activity, but also a lack of habituation to new surroundings. The study authors noted that thisnonhabituating behavior profile (i.e., decreased activity early in the test period and increased activity latein the test period) has also been reported in adult mice neonatally exposed to certain PCB congeners. Several other 1-day exposure studies have reported similar findings in rats and mice following exposureto various lower-brominated PBDEs. Decreased spontaneous activity and/or impaired habituation wasobserved in rats exposed to BDE 99 at 8 mg/kg on PND 10, mice exposed to 2,2’,4,4’,5,5’-hexaBDE(BDE 153) at ≥0.45 mg/kg on PND 10, mice exposed to BDE 47 at 10.5 mg/kg on PND 10, mice exposed to the 2,2’,3,4,4’,5’,6-heptaBDE (BDE 183) at 15.2 mg/kg on PND 3, and mice exposed to the2,2’,3,4,4’,5,5’,6-octaBDE (BDE 203) at 16.8 mg/kg on PND 3 or 10 (Eriksson et al. 2001; Viberg et al. 2003a, 2005, 2006). Increased vertical activity was significantly increased at 4 months, but not at2 months, in mice exposed to BDE 47 at ≥1 mg/kg on PND 10; no changes were observed in horizontalactivity or habituation (Gee and Moser 2008). No changes in open-field behavior were observed in mice exposed to BDE 183 at 15.2 mg/kg or 2,2’,3,3’,4,4’,5,5’,6-nonaBDE (BDE 206) at 18.5 on PND 10 (Viberg et al. 2006).Additional neurobehavioral changes observed in the studies described above included learning and memory impairments in the Morris water maze or radial arm maze in mice exposed to BDE 99 at0.8 mg/kg on PND 10, mice exposed to BDE 153 at ≥0.9 mg/kg on PND 10, and mice exposed to BDE 203 at 16.8 mg/kg on PND 10, and in rats exposed to BDE 47 at ≥1 mg/kg on PND 10 (Fischer et al.2008; He et al. 2009, 2011; Viberg et al. 2003a, 2006).Support for decreased serum T4 in F0 dams as a co-critical end point: Numerous studies report reduced serum T4 levels in adult, nonpregnant mice and rats following acute exposure to commercial pentaBDEmixtures (Bromkal 70, Bromkal 70-5 DE, DE-71), the commercial octaBDE mixture DE-79, or BDE 47. Significant reductions of 19–92% have been reported following gavage exposure at doses ≥10 and ≥0.8 mg/kg/day in rats and mice, respectively, for 1–14 days (Darnerud and Sinjari 1996; Fowles et al.1994; Hallgren and Darnerud 1998, 2002; Hallgren et al. 2001;Hoppe and Carey 2007; Richardson et al. 2008; Stoker et al. 2004, 2005; Zhou et al. 2001). Toxicokinetic considerations: Hydroxylated metabolites (OH-PBDEs) may be responsible for the toxic action of PBDEs. There is some evidence that the CYPs involved in metabolism of PBDEs, as well as the OH-PBDEs formed, are different in rats and humans (Erratico et al. 2011, 2012, 2013)—see Section 3.4.3 for more information. However, there is no available evidence indicating that the complement ofOH-PBDEs formed in humans is any more or less potent than the complement of OH-PBDEs formed in rats. Therefore, the current use of an animal-to-human uncertainty factor is appropriate for calculating an MRL based on effects observed in rats.Agency Contact (Chemical Manager): Hana Pohl PBDEs A-10APPENDIX AMINIMAL RISK LEVEL (MRL) WORKSHEETChemical Name: Polybrominated Diphenyl Ethers (PBDEs)[lower-brominated diphenyl ethers]CAS Numbers: 5436-43-1 (2,2’,4,4’-tetraBDE)Date: March 2017Profile Status: FinalRoute: [ ] Inhalation [X] OralDuration: [ ] Acute [X] Intermediate [ ] ChronicGraph Key: 164Species: RatMinimal Risk Level: 0.000003 (3x10-6) [X] mg/kg/day [ ] ppm [ ] mg/m3 Reference: Zhang Z, Zhang X, Sun Z, et al. 2013b. Cytochrome P450 3A1 mediates 2,2',4,4'-tetrabromodiphenyl ether-induced reduction of spermatogenesis in adult rats. PLoS ONE 8(6):e66301. http://www.plosone.org/article/info%3Adoi%2F10.1371%2Fjournal.pone.0066301. August 14, 2014. Experimental design: Male rats (20/group) were administered 2,2’,4,4’-tetrabromodiphenyl ether(BDE 47; ≥98.7%) at 0, 0.001, 0.03, or 1 mg/kg/day via gavage in corn oil 6 days/week for 8 weeks. Twenty-four hours after the final BDE 47 treatment, rats were sacrificed. Testes were fixed forhistological analysis and labeling of apoptotic cells or prepared for analysis of sperm production. Dailysperm production was estimated by dividing the total number of mature spermatids per testis by 6.1 (i.e., the days of the seminiferous cycle that the spermatids are present in the seminiferous epithelium).Testicular samples were examined for ROS and mRNA expression of apoptosis related proteins (ser15,ser473, p53, PTEN, AKT, BAD, caspase 3, FAS, FASL). Serum levels of E2, FSH, LH, and testosteronewere measured.Effects noted in study and corresponding doses: Histological examination of the testes showed a significant increase in the number of multinucleated giant cells (arising from spermatocytes that aborted meiosis) at ≥0.03 mg/kg/day and abundant vacuolar spaces in the seminiferous epithelium at 1 mg/kg/day(quantitative data not reported). Additionally, the number of apoptotic cells was significantly increased by 1.9- and 3-fold in the testes of rats from the 0.03 and 1 mg/kg/day groups, respectively, and the mRNAlevels of several apoptosis genes were elevated in a dose-related manner. Daily sperm production wassignificantly decreased by 23% in the 1 mg/kg/day group, compared with controls. Serum testosterone was significantly decreased by ~34, 53, and 62% in the 0.001, 0.03, and 1 mg/kg/day groups, respectively, compared with controls. No exposure-related changes were observed in serum E2, FSH, orLH levels. Testicular ROS levels were significantly elevated at 1 mg/kg/day, compared with controls. Dose and end point used for MRL derivation: 0.001 mg/kg/day[ ] NOAEL [X ] LOAEL [ ] BMDL1SDBMD modeling was performed on the serum testosterone data to assess suitability of this approach fordetermining the POD. Since testosterone data were presented graphically, GrabIt! software was used toextract the means and standard deviations. The data are shown in Table A-1. PBDEs A-11APPENDIX ATable A-1. Digitally Extracted Serum Testosterone Levels Following Exposure toBDE 47 (Zhang et al. 2013b)Dose (mg/kg) Mean testosterone level (ng/mL) Standard deviation (ng/mL)0 9.8 0.80.001 6.5 1.60.03 4.6 1.11 3.7 1.5Modeling was performed using the reference benchmark response (BMR) of 1 standard deviation changefrom the mean (1SD). Results are shown in Table A-2 and Figure A-1.Table A-2. Modeling Results for Decreased Serum Testosterone Levels Following Exposure to BDE 47 (Zhang et al. 2013b)Test for Scaled residualsc Modelsignificantdifferencep-valuea Variancep-valueb Meansp-valueb Dose belowBMDDose above BMD AICOverall largestBMD1SD (mg/kg)BMDL1SD (mg/kg)Constant varianceExponential(model 2)d <0.0001 0.14 <0.0001 -3.34 0.20 3.96 108.73 0.54 0.32Exponential(model 3)d <0.0001 0.14 <0.0001 -3.34 0.20 3.96 108.73 0.54 0.32Exponential(model 4)d <0.0001 0.14 0.11 0.00 0.00 1.13 66.79 0.0003 0.0002Exponential(model 5)d <0.0001 0.14 N/A 0.00 0.00 1.13 68.79 0.0003 0.0002Hilld <0.0001 0.14 0.18 0.02 -0.10 1.00 66.09 0.0002 0.0001Lineare <0.0001 0.14 <0.0001 -3.34 0.10 3.99 109.09 0.65 0.45Polynomial(2-degree)e <0.0001 0.14 <0.0001 -3.34 0.10 3.99 109.09 0.65 0.45Polynomial(3-degree)e <0.0001 0.14 <0.0001 -3.34 0.10 3.99 109.09 0.65 0.45Powerd,e <0.0001 0.14 <0.0001 -3.34 0.10 3.99 109.09 0.65 0.45aValues >0.05 fail to meet conventional goodness-of-fit criteria.bValues <0.10 fail to meet conventional goodness-of-fit criteria.cScaled residuals at doses immediately below and above the BMD; also the largest residual at any dose.dPower restricted to ≥1.eCoefficients restricted to be negative.AIC = Akaike Information Criterion; BDE = brominated diphenyl ether; BMD = maximum likelihood estimate of the exposure concentration associated with the selected benchmark response; BMDL = 95% lower confidence limit on the BMD (subscripts denote benchmark response: i.e., 10 = exposure concentration associated with 10% extra risk);NA = not applicable (BMDL computation failed or the BMD was higher than the highest dose tested); SD = standard deviation PBDEs A-12APPENDIX AHill Model, with BMR of 1 Std. Dev. for the BMD and 0.95 Lower Confidence Limit for the BMDL 11 10 9 8 7 6 5 4 3 2 dose 10:58 09/18 2014 Figure A-1. Fit of Hill Model to Data for Decreased Serum Testosterone Levels Following Exposure to BDE 47 (Zhang et al. 2013b)Goodness-of-fit statistics indicate inadequate fit to the data for all models except the Exponential 4 andHill models, which were considered unsuitable for use in MRL derivation because they did not providereliable information about the shape of the dose-response curve. For example, using the BMR of 1SD, the BMD, which should be within the range of the data points for best model performance, is a full orderof magnitude lower than the lowest dose used in the study. When an alternate BMR of 50% change fromthe mean (50RD) was used in order to get the BMD within the range of observation, the BMDLcalculation failed (data not shown). The observed instability in the BMD and BMDL calculationsindicates that the model is not suitable for use in MRL derivation.In the absence of a suitable model, the minimal LOAEL of 0.001 mg/kg/day for decreased serumtestosterone was chosen as the POD for MRL derivation; no NOAEL was identified. The change intestosterone is considered a minimal LOAEL because it is unclear if the magnitude of change represents a biologically adverse effect; however, this statistically significant reduction in serum testosterone is considered an early indication of damage to the male reproductive system, considering the additionaleffects observed at ≥0.03 mg/kg/day (histological lesions in testes, sperm effects). Mean Response BMDBMDLHill 0 0.2 0.4 0.6 0.8 1 PBDEs A-13APPENDIX AUncertainty factors used in MRL derivation:[X] 3 for use of a minimal LOAEL[X] 10 for extrapolation from animals to humans[X] 10 for human variabilityWas a conversion factor used from ppm in food or water to a mg/body weight dose? Not applicable(gavage study).Was a conversion used from intermittent to continuous exposure? Not applicable.If an inhalation study in animals, list conversion factors used in determining human equivalent dose: Not applicable.Other additional studies or pertinent information that lend support to this MRL: One additional rat studyand a mouse study reported histopathological changes in the testes following intermediate-duration exposure to tetraBDE at ≥0.03 mg/kg/day; neither study evaluated serum testosterone levels (Huang et al. 2015; Wang et al. 2013). In the rat study, a NOAEL of 0.001 mg/kg/day and a LOAEL of0.03 mg/kg/day were identified for increased epithelial thickness and spermatocyte apoptosis in the testes of males exposed to BDE 47 for 8 weeks via gavage (Huang et al. 2015). In the mouse study, a NOAELof 0.0015 mg/kg/day and a LOAEL of 0.045 mg/kg/day were identified for germ cell loss and increasedapoptosis in the testes of males exposed to BDE 47 for 30 days via gavage (Wang et al. 2013). Testis sections in control and 0.0015 mg/kg/day groups were normal. In the 0.045, 0.15, and 30 mg/kg/day groups, “some” seminiferous tubules exhibited complete germ cell loss and had a Sertoli cell-onlyphenotype (no incidence data reported). No exposure-related changes were observed in Leydig cells.The TUNEL assay showed a significant, dose-related increase in the number of apoptotic cells.Quantitative data were not reported; however, from the qualitative figures, it appears that apoptotic cellswere observed at doses ≥0.045 mg/kg/day. No other study evaluated testicular histopathology or serum testosterone levels in male laboratory animals following exposure to tetraBDE (BDE47). Following intermediate exposure to other congeners, no changes in testicular histology were observed in rats exposed to commercial pentaBDE mixtures(Bromkal 70-5 DE; DE-71) at gavage doses up to 250 mg/kg/day for 15–28 days (Becker et al. 2012;Oberg et al. 2010), commercial penta- or octaBDE mixtures (DE-71, unspecified octa mixture) at dietarydoses up to 750 mg/kg/day for 28–90 days (IRDC 1976, 1977; WIL Research Laboratories 1984), or adietary PBDE mixture containing 52.1% pentaBDE (DE-71), 44.2% decaBDE (BDE 209), and 0.4%octaBDE (DE-79) at doses up to 20 mg/kg/day for 70 days (Ernest et al. 2012). However, testicularapoptosis was not evaluated in any of these studies. Serum testosterone was significantly decreased by40–45% in rats exposed once to BDE 99 at 0.06 or 1.2 mg/kg (Alonso et al. 2010). Other studiesevaluating serum testosterone levels after intermediate-duration exposure to lower-brominated PBDEsmixtures (DE-71, dietary PBDE mixture described above) did not report exposure-related decreases (Becker et al. 2012; Ernest et al. 2012; Stoker et al. 2005). These data suggest that individual congeners (BDE 47, BDE 99) may have a greater capacity to alter serum testosterone levels than PBDE mixtures.One-generation studies of the BDE 47 congener reported developmental effects at ≥0.03 mg/kg/day, including:• Impaired spatial learning in the Barnes maze in PNW 8 offspring of mouse dams fed cornflakes dosed with BDE 47 from pre-mating day 28 through PND 21 (Koenig et al. 2012). PBDEs A-14APPENDIX A• Decreased center-field activity in an open field (indicating increased anxiety) in PND 60 femaleoffspring from mouse dams fed cornflakes dosed with BDE 47 from pre-mating day 28 through PND 21 (Ta et al. 2011).• Decreased pre-weaning weight, decreased pup vocalizations on PNDs 8–10, and decreased sociability on PND 72 in female offspring of mouse dams exposed to BDE 47 via gavage frompre-mating day 28 through PND 21 (Woods et al. 2012)Hydroxylated metabolites (OH-PBDEs) may be responsible for the toxic action of PBDEs. There is some evidence that the CYPs involved in metabolism of PBDEs, as well as the OH-PBDEs formed, are different in rats and humans (Erratico et al. 2011, 2012, 2013)—see Section 3.4.3 for more information. However, there is no available evidence indicating that the complement of OH-PBDEs formed in humansis any more or less potent than the complement of OH-PBDEs formed in rats. Therefore, the current use of an animal-to-human uncertainty factor is appropriate for calculating an MRL based on effects observedin rats.Agency Contact (Chemical Manager): Hana Pohl PBDEs A-15APPENDIX AMINIMAL RISK LEVEL (MRL) WORKSHEETChemical Name: Polybrominated Diphenyl Ethers (PBDEs)[decabromodiphenyl ether (decaBDE)]CAS Numbers: 1163-19-5 Date: February 2017Profile Status: FinalRoute: [ ] Inhalation [X] OralDuration: [X] Acute [ ] Intermediate [ ] ChronicGraph Key: 12Species: MouseMinimal Risk Level: 0.01 [X] mg/kg/day [ ] ppm [ ] mg/m3 References: Johansson N, Viberg H, Fredriksson A, et al. 2008. Neonatal exposure to deca-brominated diphenyl either (PBDE 209) causes dose-response changes in spontaneous behavior and cholinergic susceptibility in adult mice. Neurotoxicology 29:911-919.Buratovic S, Viberg H, Fredriksson A, et al. 2014. Developmental exposure to the polybrominated diphenyl ether PBDE 209: Neurobehavioural and neuroprotein analysis in adult male and female mice. Environ Toxicol Pharmacol 38(2):570-585.Experimental design: In the first study (Johansson et al. 2008), neonatal male mice (3–4 litters/group)were given single doses of 2,2’,3,3’,4,4’,5,5’,6,6’-decaBDE (BDE 209, 98% purity) at 0, 1.34, 2.22, 13.4, or 20.1 mg/kg via gavage in a 20% fat emulsion vehicle (1:10 mixture egg lecithin and peanut oil) on PND 3. Mice were observed for clinical signs of toxicity and body weight was measured at PND 3 and PNW 4. Spontaneous motor behavior (locomotion, rearing, total activity) was evaluated in an open field at 2 months (10 mice/group) and 4 months (16 mice/group). Motor activity was measured during a 60minute period, divided into three 20-minute intervals. Nicotine-induced behavior was evaluated at 4months following single subcutaneous injections of 80 µg nicotine/kg (8/group) or 10 mL 0.9% NaCl/kg(8/group). Anxiety was assessed at 4 months using the elevated plus maze.In the second study (Buratovic et al. 2014), neonatal male mice (6 litters/group; 31–40 males and 23– 34 females per group) were administered 2,2’,3,3’,4,4’,5,5’,6,6’-decaBDE (BDE 209, >95% purity) at doses of 0, 1.34, 5.76, or 13.4 mg/kg via gavage in a 20% fat emulsion vehicle (1:10 mixture egg lecithinand peanut oil) on PND 3. Mice were observed for clinical signs of toxicity and body weight changesthroughout the study (no further details were provided). Spontaneous motor behavior (locomotion, rearing, total activity) was evaluated in an open field at 2 months (18/sex/group). Motor activity wasmeasured during a 60-minute period, divided into three 20-minute intervals. Directly after spontaneousmotor evaluation, 9/sex/group were injected with a cholinergic agent (0.25 mg/kg paraoxon in males, 80 µg/kg nicotine in females), while the other 9/sex/group were injected with 0.9% saline, for evaluation of cholinergic-induced locomotion. At 4 months, spontaneous behavior was assessed again in the saline-injected animals only (9 males/group at all doses and 9 females/group in the control and high-dose group only). Learning and memory was assessed using the Morris water maze at 5 and 7 months in 13– 15 males from the 0, 5.76, and 13.4 mg/kg groups only (the same mice were evaluated at each time point). Male and female mice were sacrificed at 7 months. The cerebral cortex and hippocampus from controland high-dose males and females were removed and processed for neuroprotein analysis using Westernblot. PBDEs A-16APPENDIX AEffects noted in study and corresponding doses:Study 1 (Johansson et al. (2008): No clinical signs of toxicity or body weight effects were observed. At2 months, significantly decreased locomotion, rearing, and total activity were observed during the first20-minute interval of the open field assessment in mice exposed to ≥2.22 mg/kg, compared with controls. However, during the third 20-minute interval, when activity should decrease due to habituation, locomotion, rearing, and total activity were significantly increased in mice exposed to ≥13.4 mg/kg. None of the end points measured were significantly altered at 1.34 mg/kg. At 4 months, significantlydecreased locomotion, rearing, and total activity were observed during the first interval of the open fieldassessment in mice exposed to ≥2.22 mg/kg, compared with controls. During the third interval, significantly increased locomotion, rearing, and total activity were observed in mice exposed to ≥2.22 mg/kg. Additionally, total activity, but not rearing or locomotion, was significantly decreased during the first 20-mintue interval in the 1.34 mg/kg group; no significant changes were observed duringthe third interval in the 1.34 mg/kg group. Statistical analysis shows that habituation ability declined in mice exposed to ≥2.22 mg/kg from 2 to 4 months of age. At 4 months, nicotine exposure causedsignificantly decreased activity during the first interval in mice exposed to ≥13.4 mg/kg, compared with saline-injected mice from the same decaBDE exposure group. This finding is the opposite of theexpected increase in activity due to nicotine exposure, which was observed in controls and lower dosedecaBDE groups. During third interval, mice exposed to ≥13.4 mg/kg and nicotine showed impaired habituation. No exposure-related effects were observed in the elevated plus maze assessment.Study 2 (Buratovic et al. 2014): No clinical signs of toxicity or body weight effects were observed. In spontaneous activity assessment, a dose-related decrease in locomotion, rearing, and total activity wasobserved during the first 20 minutes of open field testing in a novel environment at 2 months. Decreaseswere significant at all doses tested in both sexes; however, findings were only dose-related for total activity. However, during the third 20-minute interval, when activity should decrease due to habituation, locomotion, rearing, and total activity were significantly increased in males and females at ≥5.76 mg/kg.At 2 months, cholinergic agents caused decreased activity during the first interval in mice exposed to ≥5.76 mg/kg, compared with saline-injected mice from the same decaBDE exposure group. This findingis the opposite of the expected increase in activity due to paraoxon or nicotine exposure, which wasobserved in controls and low-dose decaBDE groups. During the third interval, mice exposed to≥5.76 mg/kg and cholinergic agent showed impaired habituation. At 4 months, total activity during thefirst 20 minutes was still significantly decreased at all doses in males, and locomotion and rearing were significantly decreased in males in the mid- and high-dose groups only; all three parameters were significantly decreased in high-dose females (other doses not evaluated). All three parameters were significantly increased in high-dose males and females during the third 20-minute period, indicatingdecreased habituation; locomotion and rearing were also slightly, but significantly, increased in mid-dose males. In the Morris water maze, initial learning was comparable between exposed and control mice at5 and 7 months. However, latencies to find the escape platform during the reversal learning phase(learning to find the escape platform in a new location after initial training) were significantly longer inmid- and high-dose males at 5 and 7 months (other exposure groups not assessed). After sacrifice,significant increases in protein levels of CaMKII, Gap-43, and Tau were observed in the cortex and hippocampus in male mice and increased levels of Tau were observed in the cortex and hippocampus offemale mice. No changes in synaptophysin were observed.Dose and end point used for MRL derivation: 1.34 mg/kg[X] NOAEL [ ] LOAELIn the first study (Johansson et al. 2008), a NOAEL of 1.34 mg/kg and a LOAEL of 2.22 mg/kg weredetermined for the nonhabituating profile (i.e., decreased activity early in the test period and increased PBDEs A-17APPENDIX Aactivity late in the test period). The singular finding of decreased total activity during the first 20-minuteinterval at 4 months in the 1.34 mg/kg group was not considered sufficient to establish a LOAEL of1.34 mg/kg. The nonhabituating profile, which is a common effect observed with developmental PBDEexposure (Eriksson et al. 2002b, 2006; Fischer et al. 2008; Sand et al. 2004; Viberg et al. 2002, 2003a, 2004a, 2004b), was considered to be a stronger basis for a NOAEL/LOAEL determination. BMDmodeling was performed on the habituation ratio (activity during the last 20-minute interval/activityduring the first 20-minute interval) at 2 and 4 months reported by Johansson et al. (2008) to assess suitability of this approach for determining the POD. However, none of the models provided an adequatefit with constant or nonconstant variance.In the second study (Buratovic et al. 2014), a NOAEL of 1.34 mg/kg and a LOAEL of 5.76 mg/kg weredetermined for the nonhabituating profile (i.e., decreased activity early in the test period and increasedactivity late in the test period). Similar to the Johansson et al. (2008) study, the finding of decreased totalactivity during the first 20-minute interval at 2 and 4 months in the 1.34 mg/kg group was not considered sufficient to establish a LOAEL of 1.34 mg/kg. The nonhabituating profile was considered to be astronger basis for a NOAEL/LOAEL determination, and additional neurological effects (impairedlearning, altered response to cholinergic agents) support a LOAEL of 5.76 mg/kg. The quantitativehabituation ratio was not reported by Buratovic et al. (2014); therefore, BMD modeling was notperformed for this study.Uncertainty factors used in MRL derivation:[X] 10 for extrapolation from animals to humans[X] 10 for human variabilityWas a conversion factor used from ppm in food or water to a mg/body weight dose? Not applicable(gavage study).Was a conversion used from intermittent to continuous exposure? Not applicable (single exposure study).If an inhalation study in animals, list conversion factors used in determining human equivalent dose: Not applicable.Other additional studies or pertinent information that lend support to this MRL: In a similar study, decreased spontaneous activity and impaired habituation were also observed in 2- and 6-month-old miceexposed to BDE 209 at doses ≥2.22 mg/kg on PND 3 (lowest dose tested) (Viberg et al. 2003b). Theseeffects were not observed if exposure was on PND 10 or 19 at doses up to 20.1 mg/kg (Viberg et al. 2003b). Additionally, decreased spontaneous activity was observed in 2-month-old rats followingexposure to BDE 209 doses ≥6.7 mg/kg on PND 3 (lowest dose tested) (Viberg et al. 2007). At20.1 mg/kg, impaired habituation and decreased nicotine-induced behavior were also observed. This nonhabituating behavior profile (i.e., decreased activity early in the test period and increased activity latein the test period) is consistent with neurobehavioral alterations observed following early postnatalexposure to lower-brominated PBDEs and has been reported in adult mice neonatally exposed to certain PCB congeners (see Acute MRL Worksheet for lower-brominated PBDEs for more details).Additional neurodevelopmental effects observed in mice following acute exposure to BDE 209 from PND 2 to 15 at 20 mg/kg/day via micropipette include delayed ontogeny of reflexes, increased locomotion in males at PND 70, and learning impairment and impulsivity at 16 months, but not at3 months (Rice et al. 2007, 2009). In rats, impaired learning was observed in Morris water maze inPND 25 rat offspring of dams exposed to BDE 209 from GD 1 to 14 at doses ≥30 mg/kg/day via gavage(Chen et al. 2014). PBDEs A-18APPENDIX AHydroxylated metabolites (OH-PBDEs) may be responsible for the toxic action of PBDEs. There is some evidence that the CYPs involved in metabolism of PBDEs, as well as the OH-PBDEs formed, are different in rats and humans (Erratico et al. 2011, 2012, 2013)—see Section 3.4.3 for more information. However, there is no available evidence indicating that the complement of OH-PBDEs formed in humansis any more or less potent than the complement of OH-PBDEs formed in rats. Therefore, the current use of an animal-to-human uncertainty factor is appropriate for calculating an MRL based on effects observedin rats.Agency Contact (Chemical Manager): Hana Pohl PBDEs A-19APPENDIX AMINIMAL RISK LEVEL (MRL) WORKSHEETChemical Name: Polybrominated Diphenyl Ethers (PBDEs)[decabromodiphenyl ether (decaBDE)]CAS Numbers: 1163-19-5 Date: March 2017Profile Status: FinalRoute: [ ] Inhalation [X] OralDuration: [ ] Acute [X] Intermediate [ ] ChronicGraph Key: 29Species: RatMinimal Risk Level: 0.0002 (2x10-4) [X] mg/kg/day [ ] ppm [ ] mg/m3 Reference: Zhang Z, Sun ZZ, Xiao X, et al. 2013a. Mechanism of BDE 209-induced impaired glucosehomeostasis based on gene microarray analysis of adult rat liver. Arch Toxicol 87(8):1557-1567.Experimental design: Adult male rats were administered 2,2′,3,3′,4,4′,5,5′,6,6′-decabromodiphenyl ether(BDE 209) at 0, 0.05, 1, or 20 mg/kg/day daily via gavage in corn oil for 8 weeks. Rats were observed for clinical signs of toxicity and body weights were measured every 3 days. Rats were fasted for 24 hoursafter the final gavage treatment, and then sacrificed. Body weights and heart, spleen, lung, kidney, and liver weights were recorded. Blood was collected for clinical chemistry analysis (serum total cholesterol,triglycerides, glucose, insulin, and TNF-α) and determination of plasma markers of oxidative stress(MDE, GSH, and SOD). Liver samples from three rats in the control and low-dose (0.05 mg/kg/day)groups were collected for microarray analysis (Affymetrix GeneChip), and gene ontogeny category,pathway, gene-act-network and gene co-expression analyses were conducted. Quantitative real-time-PCR was performed to quantitate gene expression to validate the gene expression data obtained frommicroarray analysis.Effects noted in study and corresponding doses: No clinical signs of toxicity or body weight effects were observed. The relative liver weight was significantly decreased at 1 and 20 mg/kg/day by 9% (absolute liver weights were not reported). No changes were observed in relative weights of heart, spleen, lung, orkidney. No exposure-related changes were reported in serum cholesterol or triglyceride levels. Serumglucose levels were significantly increased by 12, 18, and 21% in 0.05, 1, and 20 mg/kg/day groups, compared with controls. Serum insulin was significantly decreased by 50–60% at 1 and 20 mg/kg/day. Subsequent to this finding, the pancreas was evaluated histologically. Consistent with the insulin findings, morphological changes at 1 and 20 mg/kg/day included blurred boundaries among pancreatic islet cells (quantitative data not reported). Plasma SOD activity was significantly decreased in allexposed groups and plasma GSH was significantly decreased at 1 and 20 mg/kg/day. Serum TNF-α was significantly increased at 1 and 20 mg/kg/day.BDE 209 induced 1,257 liver gene transcript changes, and 18 canonical pathways were significantlyenriched. Four of them were involved in immune diseases, including autoimmune thyroid disease, graftversus-host disease, allograft rejection, and T1DM. Subsequently, gene act network and genecoexpression network found that some major histocompatibility complex molecules and TNF-α were involved in the T1DM pathway.Dose and end point used for MRL derivation: 0.05 mg/kg/day[ ] NOAEL [X] LOAEL [ ] BMDL1SD PBDEs A-20APPENDIX ABMD modeling was performed on the serum glucose data to assess suitability of this approach fordetermining the POD. Modeling was performed using the reference BMR of one standard deviationchange from the mean (1SD), as well as an alternate BMR of 20% change from the mean (RD20). Thealternate BMR of RD20 was identified by the reference value range for rat glucose levels, which varies~20% around the mean (reference mean [range] = 118.1 mg/dL [77–141 mg/dL]; Charles RiverLaboratories 1998). Results are shown in Table A-3 and Figure A-2.Table A-3. Modeling Results for Increased Serum Glucose Levels Following Exposure to BDE 209 (Zhang et al. 2013a)Test for Scaled residualsc Modelsignificantdifferencep-valuea Variancep-valueb Meansp-valueb Dose belowBMDDose above BMDOverall largest AICBMD1SD (mg/kg)BMDL1SD (mg/kg)Constant varianceExponential(model 2)d 0.0006 0.71 0.0009 2.09 -0.10 -2.66 46.79 18.09 11.79Exponential(model 3)d 0.0006 0.71 0.0009 2.09 -0.10 -2.66 46.79 18.09 11.79Exponential(model 4)d 0.0006 0.71 0.33 0.00 0.00 0.68 35.65 0.04 0.01Exponential(model 5)d 0.0006 0.71 N/A 0.00 0.00 0.68 37.65 0.04 0.01Hilld,e 0.0006 0.71 0.41 -0.01 0.08 -0.61 35.37 0.03 0.006Linearf 0.0006 0.71 0.0009 2.09 -0.11 -2.65 46.75 17.89 11.35Polynomial(2-degree)f 0.0006 0.71 0.0009 2.09 -0.11 -2.65 46.75 17.89 11.35Polynomial(3-degree)f 0.0006 0.71 0.0009 2.09 -0.11 -2.65 46.75 17.89 11.35Powerd,e 0.0006 0.71 0.0009 2.09 -0.11 -2.65 46.75 17.89 11.35BMDRD20(mg/kg)BMDLRD20(mg/kg)Hilld,f 0.0006 0.71 0.41 0.54 NA -0.61 35.37 21.84 0.05aValues >0.05 fail to meet conventional goodness-of-fit criteria.bValues <0.10 fail to meet conventional goodness-of-fit criteria.cScaled residuals at doses immediately below and above the BMD; also the largest residual at any dose.dPower restricted to ≥1.eSelected model. With constant variance model applied, the only models that provided adequate fit to the meanswere the Exponential (model 4) and the Hill models. BMDLs for models providing adequate fit were sufficiently close(differed by <2–3-fold), so the model with the lowest AIC was selected (Hill model).fCoefficients restricted to be positive.AIC = Akaike Information Criterion; BDE = brominated diphenyl ether; BMD = maximum likelihood estimate of the exposure concentration associated with the selected benchmark response; BMDL = 95% lower confidence limit on the BMD (subscripts denote benchmark response: i.e., 10 = exposure concentration associated with 10% extra risk);NA = not applicable (BMDL computation failed or the BMD was higher than the highest dose tested); RD = relative deviation; SD = standard deviation PBDEs A-21APPENDIX AHill Model, with BMR of 1 Std. Dev. for the BMD and 0.95 Lower Confidence Limit for the BMDL Mean Response 11.5 11 10.5 10 9.5 9 8.5 BMDBMDLHill 0 5 10 15 20 dose 10:00 09/18 2014 Figure A-2. Fit of Hill Model to Data for Increased Serum Glucose Levels Following Exposure to DecaBDE (Zhang et al. 2013a)Goodness-of-fit statistics indicate inadequate fit to the data for all models except the Exponential 4 and Hill models, which were considered unsuitable for use in MRL derivation because they did not providereliable information about the shape of the dose-response curve. For example, using the reference BMRof 1SD change from the mean, the ratio of BMD:BMDL is 5 for the Hill model (0.03/0.006) and 4 for theExponential Model 4 (0.04/0.01). These values are quite high and suggest that the data do not permitaccurate estimation of the BMDL. Using the BMR of RD20 for the Hill model, the ratio was much higher still (21.84/0.05=437). The fact that this ratio changes so much with BMR underscores the instability in the BMDL estimates using this model.In the absence of a suitable model, the minimal LOAEL of 0.05 mg/kg/day based on a 12% increase in serum glucose was chosen as the POD for MRL derivation. The change in glucose is considered aminimal LOAEL because it is unclear if the magnitude of change represents a biologically adverse effect;however, the increase in serum glucose is considered to be part of a spectrum of effects indicative ofaltered insulin homeostasis and toxicity to the pancreas, including decreased serum insulin and morphological changes in pancreatic islet cells observed at ≥1 mg/kg/day, following BDE 209 exposure. PBDEs A-22APPENDIX AUncertainty factors used in MRL derivation:[X] 3 for use of a minimal LOAEL[X] 10 for extrapolation from animals to humans[X] 10 for human variabilityWas a conversion factor used from ppm in food or water to a mg/body weight dose? Not applicable(gavage study).Was a conversion used from intermittent to continuous exposure? Not applicable (doses administereddaily).If an inhalation study in animals, list conversion factors used in determining human equivalent dose: Not applicable.Other additional studies or pertinent information that lend support to this MRL: The association between PBDE-exposure and diabetes has been evaluated in a few human studies. An analysis of cross-sectionalNHANES data showed a significant increase in risk of diabetes associated with serum levels of BDE 153(but not BDE 28, BDE 47, BDE 99, or BDE 100; BDE 209 was not assessed), although the risk washigher with exposure to 50–75th percentile BDE 153 levels than >75th percentile BDE 153 levels (Lim etal. 2008). Serum BDE 153 concentrations (but not BDE 28, BDE 47, BDE 85, BDE 99, BDE 100, orBDE 154) were also shown to be significantly associated with increased odds of developing gestationaldiabetes in a cohort of 258 pregnant women; again, BDE 209 was not assessed (Smarr et al. 2016). However, other cross-sectional and prospective studies found no relationship between serum PBDEconcentrations and diabetes in an adult cohort from Wisconsin (Turyk et al. 2015), an elderly cohort in Finland (Airaksinen et al. 2011), or an elderly cohort in Sweden (Lee et al. 2011).Only one other animal study evaluated the pancreas following decaBDE exposure. In rats exposed to BDE 209 via gavage for 28 days at doses of 0, 0.27, 0.82, 2.47, 7.4, 22.2, 66.7, or 200 mg/kg/day, slightor moderate insulitis was observed in the Langerhan’s islets of the “majority of samples,” but findings were not exposure-related (Van der ven et al. 2008a). Similarly, no exposure-related effects were observed for serum glucose levels (Van der ven et al. 2008a). The only other study evaluating serumglucose levels after decaBDE exposure instead reported reduced serum glucose levels in male rats exposed to 20 mg/kg/day of a dietary PBDE mixture containing 52.1% pentaBDE (DE-71), 44.2%decaBDE (BDE 209), and 0.4% octaBDE (DE-79) for 70 days (Ernest et al. 2012). The observed decreased glucose levels could be due to the pentaBDE component, as male rats exposed to thecommercial pentaBDE mixture DE-71 at doses of 0.27–200 mg/kg/day for 28 days also showed decreased glucose levels; study authors did not report the lowest dose at which glucose levels were significantly lower in male rats, but they reported a BMD10RD of 179.55 mg/kg/day and a BMDL10RD of66.7 mg/kg/day (Van der ven et al. 2008b). Other effects occurred at doses 4–40-fold higher than theobserved pancreatic and related effects:• A LOAEL of 2 mg/kg/day was identified for transient histopathological effects in the liver ofmale offspring and kidney of female offspring of rat dams exposed to BDE 209 from GD 10 to PND 21 (no NOAEL identified) (Fujimoto et al. 2011).• A LOAEL of 10 mg/kg/day was identified for hepatocytic swelling in the liver, vacuolization in the interstitial cells of testes, and sperm damage in PND 71 male offspring of mouse damsexposed to BDE 209 from GD 0 to 17 (no NOAEL identified) (Tseng et al. 2008, 2013). PBDEs A-23APPENDIX A• A LOAEL of 20 mg/kg/day was identified for decreased anxiety in mice treated with BDE 209 by daily gavage for 15 days (no NOAEL identified) (Heredia et al. 2012).• A LOAEL of 20.1 mg/kg/day was identified for altered hippocampal electrophysiology in ratsexposed to BDE 209 from GD 1 to PND 41, PNDs 1–21, or PNDs 22–41 (no NOAEL identified)(Xing et al. 2009).Hydroxylated metabolites (OH-PBDEs) may be responsible for the toxic action of PBDEs. There is some evidence that the CYPs involved in metabolism of PBDEs, as well as the OH-PBDEs formed, are different in rats and humans (Erratico et al. 2011, 2012, 2013)—see Section 3.4.3 for more information. However, there is no available evidence indicating that the complement of OH-PBDEs formed in humansis any more or less potent than the complement of OH-PBDEs formed in rats. Therefore, the current use of an animal-to-human uncertainty factor is appropriate for calculating an MRL based on effects observedin rats.Agency Contact (Chemical Manager): Hana Pohl PBDEs A-24APPENDIX AThis page is intentionally blank. PBDEs B-1APPENDIX B. USER'S GUIDEChapter 1Public Health StatementThis chapter of the profile is a health effects summary written in non-technical language. Its intendedaudience is the general public, especially people living in the vicinity of a hazardous waste site orchemical release. If the Public Health Statement were removed from the rest of the document, it wouldstill communicate to the lay public essential information about the chemical.The major headings in the Public Health Statement are useful to find specific topics of concern. The topics are written in a question and answer format. The answer to each question includes a sentence thatwill direct the reader to chapters in the profile that will provide more information on the given topic.Chapter 2Relevance to Public HealthThis chapter provides a health effects summary based on evaluations of existing toxicologic,epidemiologic, and toxicokinetic information. This summary is designed to present interpretive, weightof-evidence discussions for human health end points by addressing the following questions:1. What effects are known to occur in humans?2. What effects observed in animals are likely to be of concern to humans?3. What exposure conditions are likely to be of concern to humans, especially around hazardouswaste sites?The chapter covers end points in the same order that they appear within the Discussion of Health Effectsby Route of Exposure section, by route (inhalation, oral, and dermal) and within route by effect. Human data are presented first, then animal data. Both are organized by duration (acute, intermediate, chronic). In vitro data and data from parenteral routes (intramuscular, intravenous, subcutaneous, etc.) are also considered in this chapter. The carcinogenic potential of the profiled substance is qualitatively evaluated, when appropriate, usingexisting toxicokinetic, genotoxic, and carcinogenic data. ATSDR does not currently assess cancerpotency or perform cancer risk assessments. Minimal Risk Levels (MRLs) for noncancer end points (ifderived) and the end points from which they were derived are indicated and discussed.Limitations to existing scientific literature that prevent a satisfactory evaluation of the relevance to public health are identified in the Chapter 3 Data Needs section.Interpretation of Minimal Risk LevelsWhere sufficient toxicologic information is available, ATSDR has derived MRLs for inhalation and oralroutes of entry at each duration of exposure (acute, intermediate, and chronic). These MRLs are notmeant to support regulatory action, but to acquaint health professionals with exposure levels at which adverse health effects are not expected to occur in humans. PBDEs B-2APPENDIX BMRLs should help physicians and public health officials determine the safety of a community living neara chemical emission, given the concentration of a contaminant in air or the estimated daily dose in water. MRLs are based largely on toxicological studies in animals and on reports of human occupationalexposure.MRL users should be familiar with the toxicologic information on which the number is based. Chapter 2, "Relevance to Public Health," contains basic information known about the substance. Other sections such as Chapter 3 Section 3.9, "Interactions with Other Substances,” and Section 3.10, "Populations that areUnusually Susceptible" provide important supplemental information.MRL users should also understand the MRL derivation methodology. MRLs are derived using a modified version of the risk assessment methodology that the Environmental Protection Agency (EPA)provides (Barnes and Dourson 1988) to determine reference doses (RfDs) for lifetime exposure. To derive an MRL, ATSDR generally selects the most sensitive end point which, in its best judgement,represents the most sensitive human health effect for a given exposure route and duration. ATSDRcannot make this judgement or derive an MRL unless information (quantitative or qualitative) is available for all potential systemic, neurological, and developmental effects. If this information and reliablequantitative data on the chosen end point are available, ATSDR derives an MRL using the most sensitive species (when information from multiple species is available) with the highest no-observed-adverse-effectlevel (NOAEL) that does not exceed any adverse effect levels. When a NOAEL is not available, a lowest-observed-adverse-effect level (LOAEL) can be used to derive an MRL, and an uncertainty factor(UF) of 10 must be employed. Additional uncertainty factors of 10 must be used both for human variability to protect sensitive subpopulations (people who are most susceptible to the health effects caused by the substance) and for interspecies variability (extrapolation from animals to humans). In deriving an MRL, these individual uncertainty factors are multiplied together. The product is then divided into the inhalation concentration or oral dosage selected from the study. Uncertainty factors usedin developing a substance-specific MRL are provided in the footnotes of the levels of significant exposure (LSE) tables.Chapter 3Health EffectsTables and Figures for Levels of Significant Exposure (LSE)Tables and figures are used to summarize health effects and illustrate graphically levels of exposure associated with those effects. These levels cover health effects observed at increasing dose concentrations and durations, differences in response by species, MRLs to humans for noncancer endpoints, and EPA's estimated range associated with an upper- bound individual lifetime cancer risk of 1 in 10,000 to 1 in 10,000,000. Use the LSE tables and figures for a quick review of the health effects and to locate data for a specific exposure scenario. The LSE tables and figures should always be used inconjunction with the text. All entries in these tables and figures represent studies that provide reliable,quantitative estimates of NOAELs, LOAELs, or Cancer Effect Levels (CELs).The legends presented below demonstrate the application of these tables and figures. Representative examples of LSE Table 3-1 and Figure 3-1 are shown. The numbers in the left column of the legendscorrespond to the numbers in the example table and figure. PBDEs B-3APPENDIX BLEGENDSee Sample LSE Table 3-1 (page B-6)(1) Route of Exposure. One of the first considerations when reviewing the toxicity of a substance using these tables and figures should be the relevant and appropriate route of exposure. Typicallywhen sufficient data exist, three LSE tables and two LSE figures are presented in the document. The three LSE tables present data on the three principal routes of exposure, i.e., inhalation, oral, and dermal (LSE Tables 3-1, 3-2, and 3-3, respectively). LSE figures are limited to the inhalation (LSE Figure 3-1) and oral (LSE Figure 3-2) routes. Not all substances will have data on eachroute of exposure and will not, therefore, have all five of the tables and figures.(2) Exposure Period. Three exposure periods—acute (less than 15 days), intermediate (15– 364 days), and chronic (365 days or more)—are presented within each relevant route of exposure.In this example, an inhalation study of intermediate exposure duration is reported. For quickreference to health effects occurring from a known length of exposure, locate the applicableexposure period within the LSE table and figure.(3) Health Effect. The major categories of health effects included in LSE tables and figures are death, systemic, immunological, neurological, developmental, reproductive, and cancer. NOAELs and LOAELs can be reported in the tables and figures for all effects but cancer.Systemic effects are further defined in the "System" column of the LSE table (see key number18).(4) Key to Figure. Each key number in the LSE table links study information to one or more datapoints using the same key number in the corresponding LSE figure. In this example, the studyrepresented by key number 18 has been used to derive a NOAEL and a Less Serious LOAEL(also see the two "18r" data points in sample Figure 3-1).(5) Species. The test species, whether animal or human, are identified in this column. Chapter 2,"Relevance to Public Health," covers the relevance of animal data to human toxicity andSection 3.4, "Toxicokinetics," contains any available information on comparative toxicokinetics. Although NOAELs and LOAELs are species specific, the levels are extrapolated to equivalenthuman doses to derive an MRL.(6) Exposure Frequency/Duration. The duration of the study and the weekly and daily exposureregimens are provided in this column. This permits comparison of NOAELs and LOAELs fromdifferent studies. In this case (key number 18), rats were exposed to “Chemical x” via inhalationfor 6 hours/day, 5 days/week, for 13 weeks. For a more complete review of the dosing regimen,refer to the appropriate sections of the text or the original reference paper (i.e., Nitschke et al.1981).(7) System. This column further defines the systemic effects. These systems include respiratory,cardiovascular, gastrointestinal, hematological, musculoskeletal, hepatic, renal, anddermal/ocular. "Other" refers to any systemic effect (e.g., a decrease in body weight) not coveredin these systems. In the example of key number 18, one systemic effect (respiratory) was investigated.(8) NOAEL. A NOAEL is the highest exposure level at which no harmful effects were seen in the organ system studied. Key number 18 reports a NOAEL of 3 ppm for the respiratory system, which was used to derive an intermediate exposure, inhalation MRL of 0.005 ppm (see footnote "b"). PBDEs B-4APPENDIX B(9) LOAEL. A LOAEL is the lowest dose used in the study that caused a harmful health effect. LOAELs have been classified into "Less Serious" and "Serious" effects. These distinctions help readers identify the levels of exposure at which adverse health effects first appear and the gradation of effects with increasing dose. A brief description of the specific end point used to quantify the adverse effect accompanies the LOAEL. The respiratory effect reported in keynumber 18 (hyperplasia) is a Less Serious LOAEL of 10 ppm. MRLs are not derived fromSerious LOAELs.(10) Reference. The complete reference citation is given in Chapter 9 of the profile.(11) CEL. A CEL is the lowest exposure level associated with the onset of carcinogenesis inexperimental or epidemiologic studies. CELs are always considered serious effects. The LSEtables and figures do not contain NOAELs for cancer, but the text may report doses not causing measurable cancer increases.(12) Footnotes. Explanations of abbreviations or reference notes for data in the LSE tables are found in the footnotes. Footnote "b" indicates that the NOAEL of 3 ppm in key number 18 was used to derive an MRL of 0.005 ppm.LEGENDSee Sample Figure 3-1 (page B-7)LSE figures graphically illustrate the data presented in the corresponding LSE tables. Figures help thereader quickly compare health effects according to exposure concentrations for particular exposureperiods.(13) Exposure Period. The same exposure periods appear as in the LSE table. In this example, healtheffects observed within the acute and intermediate exposure periods are illustrated.(14) Health Effect. These are the categories of health effects for which reliable quantitative data exists. The same health effects appear in the LSE table.(15) Levels of Exposure. Concentrations or doses for each health effect in the LSE tables are graphically displayed in the LSE figures. Exposure concentration or dose is measured on the logscale "y" axis. Inhalation exposure is reported in mg/m3 or ppm and oral exposure is reported in mg/kg/day.(16) NOAEL. In this example, the open circle designated 18r identifies a NOAEL critical end point in the rat upon which an intermediate inhalation exposure MRL is based. The key number 18 corresponds to the entry in the LSE table. The dashed descending arrow indicates theextrapolation from the exposure level of 3 ppm (see entry 18 in the table) to the MRL of0.005 ppm (see footnote "b" in the LSE table).(17) CEL. Key number 38m is one of three studies for which CELs were derived. The diamond symbol refers to a CEL for the test species-mouse. The number 38 corresponds to the entry in theLSE table. PBDEs B-5APPENDIX B(18) Estimated Upper-Bound Human Cancer Risk Levels. This is the range associated with the upper-bound for lifetime cancer risk of 1 in 10,000 to 1 in 10,000,000. These risk levels are derived from the EPA's Human Health Assessment Group's upper-bound estimates of the slope of thecancer dose response curve at low dose levels (q1*).(19) Key to LSE Figure. The Key explains the abbreviations and symbols used in the figure. SAMPLE PBDEs B-6 APPENDIX B 1 → Table 3-1. Levels of Significant Exposure to [Chemical x] – InhalationLOAEL (effect)Exposure Less serious Serious (ppm)Key to frequency/ NOAEL(ppm)figurea Species duration System (ppm) Reference234→ INTERMEDIATE EXPOSURE5→ Systemic ↓18 Rat→CHRONIC EXPOSURECancer38 Rat39 Rat40 Mouse6↓13 wk5 d/wk6 hr/d18 mo5 d/wk7 hr/d89–104 wk5 d/wk6 hr/d79–103 wk5 d/wk6 hr/d7 8 9↓ ↓ ↓Resp 3b 10 (hyperplasia)11↓20 (CEL, multipleorgans)10 (CEL, lung tumors,nasal tumors)10 (CEL, lung tumors,hemangiosarcomas)10↓Nitschke et al. 1981Wong et al. 1982NTP 1982NTP 198212 → a The number corresponds to entries in Figure 3-1.b Used to derive an intermediate inhalation Minimal Risk Level (MRL) of 5x10-3 ppm; dose adjusted for intermittent exposure and divided by an uncertainty factor of 100 (10 for extrapolation from animal to humans, 10 for human variability). PBDEs B-7 APPENDIX B Chronic (≥ 365 days) Intermediate (15-364 days) PBDEs B-8APPENDIX BThis page is intentionally blank. PBDEs C-1APPENDIX C. ACRONYMS, ABBREVIATIONS, AND SYMBOLSACGIH American Conference of Governmental Industrial HygienistsACOEM American College of Occupational and Environmental MedicineADI acceptable daily intakeADME absorption, distribution, metabolism, and excretionAED atomic emission detectionAFID alkali flame ionization detectorAFOSH Air Force Office of Safety and HealthALT alanine aminotransferaseAML acute myeloid leukemiaAOAC Association of Official Analytical ChemistsAOEC Association of Occupational and Environmental ClinicsAP alkaline phosphataseAPHA American Public Health AssociationAST aspartate aminotransferaseatm atmosphereATSDR Agency for Toxic Substances and Disease RegistryAWQC Ambient Water Quality CriteriaBAT best available technologyBCF bioconcentration factorBEI Biological Exposure IndexBMD/C benchmark dose or benchmark concentrationBMDX dose that produces a X% change in response rate of an adverse effectBMDLX 95% lower confidence limit on the BMDX BMDS Benchmark Dose SoftwareBMR benchmark responseBSC Board of Scientific CounselorsC centigradeCAA Clean Air ActCAG Cancer Assessment Group of the U.S. Environmental Protection AgencyCAS Chemical Abstract ServicesCDC Centers for Disease Control and PreventionCEL cancer effect levelCELDS Computer-Environmental Legislative Data SystemCERCLA Comprehensive Environmental Response, Compensation, and Liability ActCFR Code of Federal RegulationsCi curieCI confidence intervalCL ceiling limit valueCLP Contract Laboratory Programcm centimeterCML chronic myeloid leukemiaCPSC Consumer Products Safety CommissionCWA Clean Water ActDHEW Department of Health, Education, and WelfareDHHS Department of Health and Human ServicesDNA deoxyribonucleic acidDOD Department of DefenseDOE Department of EnergyDOL Department of Labor PBDEs C-2APPENDIX CDOT Department of TransportationDOT/UN/ Department of Transportation/United Nations/NA/IMDG North America/Intergovernmental Maritime Dangerous Goods CodeDWEL drinking water exposure levelECD electron capture detectionECG/EKG electrocardiogramEEG electroencephalogramEEGL Emergency Exposure Guidance LevelEPA Environmental Protection AgencyF FahrenheitF1 first-filial generationFAO Food and Agricultural Organization of the United NationsFDA Food and Drug AdministrationFEMA Federal Emergency Management AgencyFIFRA Federal Insecticide, Fungicide, and Rodenticide ActFPD flame photometric detectionfpm feet per minuteFR Federal RegisterFSH follicle stimulating hormoneg gramGC gas chromatographygd gestational dayGLC gas liquid chromatographyGPC gel permeation chromatographyHPLC high-performance liquid chromatographyHRGC high resolution gas chromatographyHSDB Hazardous Substance Data BankIARC International Agency for Research on CancerIDLH immediately dangerous to life and healthILO International Labor OrganizationIRIS Integrated Risk Information SystemKd adsorption ratiokg kilogramkkg metric tonKoc organic carbon partition coefficientKow octanol-water partition coefficientL literLC liquid chromatographyLC50 lethal concentration, 50% killLCLo lethal concentration, lowLD50 lethal dose, 50% killLDLo lethal dose, lowLDH lactic dehydrogenaseLH luteinizing hormoneLOAEL lowest-observed-adverse-effect levelLSE Levels of Significant ExposureLT50 lethal time, 50% killm meterMA trans,trans-muconic acidMAL maximum allowable levelmCi millicurie PBDEs C-3APPENDIX CMCL maximum contaminant levelMCLG maximum contaminant level goalMF modifying factorMFO mixed function oxidasemg milligrammL millilitermm millimetermmHg millimeters of mercurymmol millimolemppcf millions of particles per cubic footMRL Minimal Risk LevelMS mass spectrometryNAAQS National Ambient Air Quality StandardNAS National Academy of ScienceNATICH National Air Toxics Information ClearinghouseNATO North Atlantic Treaty OrganizationNCE normochromatic erythrocytesNCEH National Center for Environmental HealthNCI National Cancer InstituteND not detectedNFPA National Fire Protection Associationng nanogramNHANES National Health and Nutrition Examination SurveyNIEHS National Institute of Environmental Health SciencesNIOSH National Institute for Occupational Safety and HealthNIOSHTIC NIOSH's Computerized Information Retrieval SystemNLM National Library of Medicinenm nanometernmol nanomoleNOAEL no-observed-adverse-effect levelNOES National Occupational Exposure SurveyNOHS National Occupational Hazard SurveyNPD nitrogen phosphorus detectionNPDES National Pollutant Discharge Elimination SystemNPL National Priorities ListNR not reportedNRC National Research CouncilNS not specifiedNSPS New Source Performance StandardsNTIS National Technical Information ServiceNTP National Toxicology ProgramODW Office of Drinking Water, EPAOERR Office of Emergency and Remedial Response, EPAOHM/TADS Oil and Hazardous Materials/Technical Assistance Data SystemOPP Office of Pesticide Programs, EPAOPPT Office of Pollution Prevention and Toxics, EPAOPPTS Office of Prevention, Pesticides and Toxic Substances, EPAOR odds ratioOSHA Occupational Safety and Health AdministrationOSW Office of Solid Waste, EPAOTS Office of Toxic Substances PBDEs C-4APPENDIX COW Office of WaterOWRS Office of Water Regulations and Standards, EPAPAH polycyclic aromatic hydrocarbonPBPD physiologically based pharmacodynamicPBPK physiologically based pharmacokinetic PCE polychromatic erythrocytesPEL permissible exposure limitpg picogramPHS Public Health ServicePID photo ionization detectorpmol picomolePMR proportionate mortality ratioppb parts per billionppm parts per millionppt parts per trillionPSNS pretreatment standards for new sourcesRBC red blood cellREL recommended exposure level/limitRfC reference concentrationRfD reference doseRNA ribonucleic acidRQ reportable quantityRTECS Registry of Toxic Effects of Chemical SubstancesSARA Superfund Amendments and Reauthorization ActSCE sister chromatid exchangeSGOT serum glutamic oxaloacetic transaminaseSGPT serum glutamic pyruvic transaminaseSIC standard industrial classificationSIM selected ion monitoringSMCL secondary maximum contaminant levelSMR standardized mortality ratioSNARL suggested no adverse response levelSPEGL Short-Term Public Emergency Guidance LevelSTEL short term exposure limitSTORET Storage and RetrievalTD50 toxic dose, 50% specific toxic effectTLV threshold limit valueTOC total organic carbonTPQ threshold planning quantityTRI Toxics Release InventoryTSCA Toxic Substances Control ActTWA time-weighted averageUF uncertainty factorU.S. United StatesUSDA United States Department of AgricultureUSGS United States Geological SurveyVOC volatile organic compoundWBC white blood cellWHO World Health Organization PBDEs C-5APPENDIX C> greater than≥ greater than or equal to= equal to< less than≤ less than or equal to% percentα alphaβ betaγ gammaδ deltaμm micrometerμg microgramq1* cancer slope factor– negative+ positive(+) weakly positive result(–) weakly negative result

Toxicological profile for polybrominated diphenyl ethers (PBDEs)

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