토양미생물의 활성을 이용한 화약류의 토양 생태독성 및 생태학적 허용농도 결정에 관한 연구

학위논문 (석사)-- 서울대학교 대학원 : 건설환경공학부, 2014. 8. 남경필.사격장 오염 토양에 대한 관심이 커지고 있으나 아직까지 화약물질에 대한 국내 법규와 정화기준이 존재하지 않아 화약물질을 정화 및 관리하기에 많은 어려움을 겪고 있다. 사격장 내 오염물질은 화약물질 뿐 아니라 중금속도 발견되나, 중금속의 경우 국내 법규 상 토양 기준이 대부분 존재하며, 중금속의 생태독성학적 영향에 관련하여 이미 많은 연구가 진행되었기 때문에 본 연구에서는 사격장 오염 부지의 주요 오염물질 중에서 화약물질을 선정하였다. 화약물질은 대표적으로 2,4,6-trinitrotoluene (TNT), hexahydro–1,3,5–trinitro–1,3,5-triazine (RDX), 1,3,5,7–tetranitro–1,3,5,7-tetrazocane (HMX) 이나 국내에서 HMX는 사용하지 않기때문에 국내 실정에 맞는 TNT와 RDX를 대상 물질로 선정하였다. 다양한 생태수용체를 대상으로 한 독성평가 결과로 도출된 독성종말점은 생태학적 토양 허용농도를 도출하는 과정에서 사용된다. 하지만 미생물활성실험을 통한 토양미생물에 대한 독성영향을 평가한 독성 자료가 부족한 실정이다. TNT의 경우 토양미생물에 대한 독성 종말점이 존재하나, 그중 일부는 미생물의 생체중량을 평가한 것으로 토양미생물 활성평가를 통한 독성자료의 수가 부족하며, RDX의 경우 토양미생물을 대상으로 한 독성영향 평가가 거의 진행되지 않았다. 따라서 본 연구에서는 토양미생물 활성평가를 통하여 차후 생태독성학적 영향 평가에 이용될 수 있는 독성자료를 제시한다. 토양미생물 활성평가를 위해 potential nitrification activity, dehydragenase activity, phosphatase activity, fluorescein diacetate activity, β-glucosidase activity, arylsulfatase activity and rhodanese activity, 7가지의 효소반응을 2주, 4주 그리고 8주의 노출기간을 두어 측정하였다. 이때 다양한 토양 특성에 따른 독성발현의 정도 변화를 알아보기 위하여 사격장토양, 임야토양, 논토양, 매립지토양, 모래, 점토, 유기물을 이용하여 실험하였다. 각 토양별, 그리고 효소반응별 실험결과에서 NOEC (No Observed Effect Concentration)을 도출한 후 하나의 효소반응에 대한 대표적 NOEC 값을 기하평균으로 계산하였다. TNT의 경우, NOEC이 45.31 (fluorescein diacetate activity)에서 55.15 (dehydrogenase activity) mg/kg에 해당하는 값을 보였고 RDX의 경우, NOEC 값이 285.9 (phosphatase activity)에서 308.9 (dehydrogenase activity) mg/kg에 해당하였다. 최근 국제적으로 생태 수용체에 대한 고려가 점차 중요해지고 있다. 토양매체의 경우 토양 생태 중심의 위해성 평가와 인체 중심의 위해성 평가 모두를 고려하여 토양 오염을 관리하는 미국, 캐나다, 네델란드, 독일 등의 국가들과 달리 우리나라에서는 아직 인체 중심의 위해성 평가만 이루어졌을 뿐 생태 수용체에 대한 고려가 없는 실정이다. 국내 서식종, 국제 표준 시험종, 국제 서식종을 포함하여 TNT와 RDX의 토양 생물에 대한 독성 지표값 350여개를 미국의 EPA(Environmental Protection Agency), 캐나다의 CCME (Canadian Council of Ministers of the Environment) 와 논문을 통해 조사하였고, 수집된 자료는 qualification 과정을 통하여 신뢰도가 높은 자료만을 선별하였다. 독성자료를 도출하는 과정에서 국제표준 독성실험법을 사용하여야 하고, 노출기간, 독성종말점, 통계학적 처리방법, 인공오염 후 실제 오염농도 그리고 그 밖의 실험 조건에 대한 정확한 명시가 되어 있어야 한다. 토양 독성 자료의 경우 토양 내 유기물 함량에 따라 독성 발현이 달라질 수 있으므로 기준 유기물함량에 맞도록 표준화 과정을 거쳤다. 조건을 만족하는 독성자료들을 기하평균을 통해 하나의 종과 하나의 독성종말점에 대한 대표값을 계산하였다. 수집된 독성자료의 quality에 따라 종민감도분포도(SSD: Species Sensitivity Distribution), Assessment Factor method (AF), 그리고 Equilibrium Partitioning method 중 하나의 방법을 선정하였다. 어떤 물질의 환경기준을 결정하는 데에는 그 물질의 독성뿐 아니라 배경농도, 가용한 정화기술, 사회경제적 고려 등이 포함되어야 하지만, 본 연구에서 도출된 TNT와 RDX에 토양 생물이 5% 영향을 받는 농도인 HC5는 생태 보호를 위한 토양의 TNT, RDX 관리기준의 기본 값으로는 사용될 수 있을 것이다. 사격장의 경우, 사람의 접근은 제한되어 있으나 주변 서식 동식물의 접근은 가능하기 때문에 토양 생태 수용체를 반드시 고려해야한다. 실제 국내 다락대 사격장과 주변지역을 조사한 결과, 고라니, 멧돼지, 다람쥐를 포함한 26 종의 포유류, 제비, 거위, 까마귀를 포함한 77 종의 조류, 부들, 갈대, 버드나무를 포함한 90 종의 식물이 발견되어 생태 수용체 고려의 필요성을 확인하였다. 본 연구에서 도출한 HC5를 PNEC (Predicted No Effect Concentration)로 사용하고 한탄강 유역에 있는 다락대 종합사격장 TNT, RDX의 토양 오염 농도를 PEC (Predicted Exposure Concentration)으로 하여 HQ（Hazard Quotient)를 도출하여 생태 위해 여부를 판단하여 보았다. 토양 오염 농도인 PEC은 다락대 종합사격장 내 실시한 incermental sampling 결과, 발견농도의 대표농도 TNT-40.59 mg/kg (95% Gamma UCL), RDX-52.97 mg/kg (95% chebyshev UCL)을 사용하였다. PNEC의 경우 본 연구에서 도출한 다락대 사격장 주변 한탄강의 수생태계를 보호할 수 있는 생태학적 허용농도인 TNT-807 mg/kg, RDX-56 mg/kg을 사용하였다. 토양에서 TNT의 HQ는 대표 농도에서 0.05, RDX의 경우 0.95이므로 1 이하 값이 산출되어 생태학적으로 위해 가능성이 없는 것으로 판단되었다. 도출한 HQ 값은 다락대 사격장 피탄지 토양에 존재하는 TNT와 RDX가 주변 수계인 한탄강의 수계생태계에 영향을 미칠 가능성이 없는 것을 의미하나, RDX의 경우 0.95이므로 추가적인 피탄지 토양에 대한 조사가 필요할 것으로 판단된다.Soil contamination with explosives at firing ranges are recently found to be influential to the surrounding ecosystem. At highly active firing ranges, it is more reasonable to manage the toxic effect of explosives on surrounding ecosystem than the direct remediation of contaminants on site. This study was performed to determine the effects of explosives-contaminated soil and water and to suggest ecologically permissible concentrations of explosives. Among explosives, 2,4,6-trinitrotoluene (TNT), hexahydro-1,3,5-trinitro-1,3,5-triazine (RDX) were selected as target pollutants. Moreover, microbial activity experiments were executed for more qualified derivation of permissible concentrations, since few microbial toxicity data of TNT and RDX were available. Toxicity data of various species were available, however, the effects of TNT and RDX on microbes were not studied widely, only few soil toxicity data of TNT and RDX on microbes were available. Therefore, the study of toxic effects of TNT and RDX on microbes is required. Soil microbial activity such as potential nitrification activity, dehydragenase activity, phosphatase activity, fluorescein diacetate activity, β-glucosidase activity, arylsulfatase activity and rhodanese activity were measured using seven different soil types. NOECs were derived to verify the toxicity of TNT and RDX on soil microbes by calculating the geometric means of NOECs using different soil types each from different enzymatic assay methods. For TNT, NOEC values varied from 45.31 (fluorescein diacetate activity) to 55.15 (dehydrogenase activity) mg/kg and for RDX, NOEC values varied from 285.9 (phosphatase activity) to 308.9 (dehydrogenase activity) mg/kg. The derivation and suggestion of permissible concentrations of TNT and RDX are required, since there is no standards of TNT and RDX in Korea. The toxicity values of various test organisms from literatures are chosen in order to derive ecological permissible concentrations. The permissible concentrations were derived using guidelines such as A Protocol for the Derivation of Environmental and Human Health Soil Quality Guidelines (CCME) and Guidance for the derivation of environmental risk limits within the framework of International and national environmental quality standards for substances in the Netherlands (RIVM). Ecologically permissible concentrations of TNT and RDX in soil are suggested as TNT-7.7 (UF=5) and 17.3 (UF=1) mg/kg, RDX-18.3 (UC=5) and 41 (UF=1) mg/kg for Dutch RIVM approach and TNT-5.6 (UF=5) and 28.1 (UF=1) mg/kg, RDX-15 (UF=5) and 75 (UF=1) mg/kg for Canadian CCME approach. Each concentrations were derived through selected procedure chosen by quality of each toxicity data sets. The permissible concentrations of TNT varies slightly less than RDX between CCME and RIVM approach. The toxicity data of RDX used for CCME approach might determined to be less qualified because EC50 values were used due to lack of EC20 values. When using this data set, RIVM approach can be determined to be more precise. For determination of environmental quality standard of each contaminants, not only toxicity but also background concentration, capability and applicability of remediation and management technology and other social/economical conditions should be considered, however ecologically permissible concentrations derived in this study can be used as screening level. HC5 values which are derived in this study were set to PNEC (Predicted No Effect Concentration), and the measured TNT and RDX concentrations in active firing range soil were set to PEC (Predicted Exposure Concentration). HQ (Hazardous Quotient) was calculated to determine the ecotoxicological risk. PEC were the representative concentration of TNT-40.59 mg/kg (95% Gamma UCL) and RDX-52.97 mg/kg (95% chebyshev UCL) using incremental sampling. Ecologically permissible firing range soil for protecting aquatic species of TNT-807 mg/kg and RDX-56 mg/kg were used for PNEC. The calculated HQ value of TNT was 0.05, and for RDX, it was 0.95 with representative concentrations indicating no ecotoxicological risk are found in this contamination site, however the derived HQ value of RDX was 0.95 which is close to one. It is determined that further investigation on RDX in active firing range is needed.1. Introduction 1 1.1 Background 1 1.2 Objectives 3 1.3 Dissertation Structure 4 2. Literature Review 5 2.1 Ecotoxicity Tests 5 2.1.1 Ecotoxicity Tests on Terrestrial Species 5 2.1.2 Ecotoxicity Tests on Aquatic Species 7 2.2 Ecotoxicity Tests on Soil Microbes 9 2.2.1 Mechanism of Enzyme Action 10 2.2.2 The Activity of Soil Microbes 11 2.2.2.1 Potential nitrification activity 11 2.2.2.2 Dehydrogenase activity 12 2.2.2.3 Phosphatase activity 13 2.2.2.4 Fluorescein diacetate hydrolytic activity 14 2.2.2.5 β-glucosidase activity 14 2.2.2.6 Arylsulfatase activity 16 2.2.2.7 Rhodanese activity 17 2.3 Ecotoxicological Risk Assessment 18 2.3.1 Probabilistic Approach 18 2.3.1.1 Species sensitivity distribution 19 2.3.2 Deterministic Approach 21 2.3.2.1 Assessment factor method 21 2.3.2.2 Equilibrium partitioning method 23 2.4 Environmental Quality Standards 25 2.4.1 Setting Soil Quality Standard 25 2.4.1.1 Canadian CCME approach 25 2.4.1.2 Dutch RIVM approach 27 2.4.2 Setting Water Quality Standard 28 3. Materials and Methods 30 3.1 Determination of Microbial Activity in Soil 30 3.1.1 Materials 30 3.1.1.1 Test soil types 30 3.1.1.2 Chemicals and reagents 32 3.1.2 Soil Microbial Activity Test Methods 33 3.1.2.1 Potential nitrification activity 33 3.1.2.2 Dehydrogenase activity 34 3.1.2.3 Phosphatase activity 34 3.1.2.4 Fluorescein diacetate hydrolytic activity 35 3.1.2.5 β-glucosidase activity 35 3.1.2.6 Arylsulfatase activity 36 3.1.2.7 Rhodanese activity 36 3.2 Derivation of Ecologically Permissible Concentrations 37 3.2.1 Toxicological Data Qualification Process 37 3.2.1.1 Data collection 37 3.2.1.2 Data screening 39 3.2.1.3 Data qualification 40 3.2.1.4 Data extrapolation type selection 41 3.2.2 Species Sensitivity Distribution (SSD) 41 3.2.2.1 The concept of SSD 41 3.2.2.2 Selection of HC5 43 3.2.3 Derivation of Ecologically Permissible Concentrations 44 3.2.3.1 Ecologically permissible soil concentration 44 3.2.3.2 Ecologically permissible water concentration 44 3.3 Analysis 45 3.3.1 Soil Microbial Activity Analysis 45 3.2.3.1 Ion Chromatography (IC) 45 3.2.3.2 UV/Vis Spectrophotometer 45 3.3.2 Statistics and Data Interpretation 45 3.3.3 TNT and RDX Analysis (HPLC) 46 4. Results and Discussion 47 4.1 Soil Microbial Activity 47 4.1.1 Soil Microbial Activity Test Results 48 4.1.1.1 Potential nitrification activity 49 4.1.1.2 Dehydrogenase activity 50 4.1.1.3 Phosphatase activity 51 4.1.1.4 Fluorescein diacetate hydrolytic activity 52 4.1.1.5 β-glucosidase activity 53 4.1.1.6 Arylsulfatase activity 54 4.1.1.7 Rhodanese activity 55 4.1.2 Toxicity Endpoint Suggestion 56 4.2 Derivation of Ecologically Permissible Concentration 57 4.2.1 Ecologically Permissible Soil Concentrations using RIVM Approach 57 4.2.2 Ecologically Permissible Soil Concentrations using CCME Approach 60 4.2.3 Comparison of Ecologically Permissible Soil Concentrations among Major Trophic Levels 63 4.3 Application in Korean Active Firing Range 66 4.3.1 Site Characterization 66 4.3.2 Derivation of Ecologically Permissible Water Concentration 68 4.3.3 Derivation of Ecologically Permissible Active Firing Range Soil Concentration by Efflux Equation 71 4.3.4 Ecotoxicological Risk Assessment at Darakdae Active Firing Range 75 5. Conclusion 77 References 81Maste

유기양이온인 tributylmethyl ammonium (TBuMA)의 소장흡수-용량증가를 초과하는 흡수증가 현상의 해석

Thesis(doctoral)--서울대학교 대학원 :제약학과 약제과학전공,2006.Docto

수체 내 수은의 이화학적 메틸화 및 디메틸화 반응의 기작 연구

학위논문 (박사)-- 서울대학교 보건대학원 : 보건학과, 2014. 8. 조경덕.Methylmercury (MeHg) is among the most widespread contaminants that pose severe health risks to humans and wildlife. In determination the levels of MeHg in aquatic environments, methylation of inorganic mercury (Hg(II)) to MeHg and demethylation of MeHg are the two most important processes in the cycling of MeHg. So, the knowledge of the efficiency of these different pathways of Hg methylation and demethylation is one of the key steps to predict MeHg concentrations in the different environmental compartments and to estimate the Hg bio-accessibility to the organisms. However, the factors that influence the competing methylation and demethylation reactions are yet insufficiently understood and little to no attempt has been made to determine end products, especially abiotic processes. The relative importance of each reaction and the resulting net effect will probably depend on the environmental conditions. Therefore, this study investigated the possible photochemical processes and mechanism of Hg demethylation and methylation in water with simulating various environmental conditions. The main objectives of this study were (1) to investigate the influence of several environmental factors and other water constituents on photo-decomposition of MeHg (Study 1), (2) to understand the mechanism of MeHg demethylation process in seawater by assessing the production of dissolved gaseous mercury (DGM) generated from MeHg photo-degradation (Study 2), and (3) to assess the possibility of various methyl donors such as acetate, malonate, dimethylsulfoxide, and litter-derived DOM for photochemical methylation of Hg(II) in aquatic systems (Study 3). For Study 1, photo-initiated decomposition of MeHg was investigated under UVA irradiation in the presence of natural water constituents including nitrate (NO3−), ferric (Fe3+), and bicarbonate (HCO3−) ions, and DOM such as humic and fulvic acid (HA and FA). MeHg degradation followed the pseudo-first-order kineticsthe rate constant increased with increasing UVA intensity ranged from 0.3 to 3.0 mW cm-2. In the presence of NO3−, Fe3+, and FA, the decomposition rate of MeHg increased significantly due to photosensitization by reactive species such as hydroxyl radical (OH•). However, the presence of HA and HCO3− ions lowered the degradation rate through a radical scavenging effect. Increasing the pH of the solution increased the degradation rate constant by enhancing the generation of OH•. Therefore, OH• play an important role in the photo-decomposition of MeHg in water, and natural constituents in water can affect the photo-decomposition of MeHg by changing radical production and inhibition. For Study 2, the photo-induced formation of dissolved gaseous mercury (DGM, Hg0) from MeHg removal was investigated. This study examined the effect of various environmental factors (i.e., light wavelength and intensity and MeHg concentration), and primary water constituents on the abiotic photo-degradation of MeHg, especially under different salinity. Photo-degradation rates of MeHg were positively correlated with the UV light intensity, implying that the attenuation of UV radiation had a significant effect on MeHg photo-degradation in water. However, a high dissolved organic carbon (DOC) concentration and salinity inhibited MeHg photo-degradation. DGM was always produced during the photo-degradation of MeHg. Photo-degradation rates of MeHg and DGM production decreased with increasing salinity, suggesting that the presence of chloride ions inhibited MeHg photo-degradation. Therefore, this study imply that MeHg in freshwater could be more rapidly demethylated than that in seawater and MeHg flowing into the lake or river would be almost removed by photo-demethylation. However, MeHg flowing to seawater would be hardly removed, which could have more chance for bioaccumulation in seawater. For Study 3, the photochemical methylation of Hg(II) using various methyl donors such as acetate, malonate, dimethylsulfoxide (DMSO), and litter-derived dissolved organic matter (LDOM) was examined. The methylation reaction via acetate was followed the pseudo-first-order kinetics for Hg(II), and the methylation ability of acetate decreased with the solution pH increased. In the Hg(II) methylation by LDOM, LDOM leaded to the production of new MeHg under not only UV irradiation but dark condition. Especially, from the results of the production new MeHg by LDOM in the microbial free and dark condition, this work suggests the possibility that the abiotic chemical reaction such as a non-dependence upon light occurs in the natural aquatic environment. In addition, for the MeHg formation of Hg(II) by DMSO in seawater, abiotic methylation reaction appeared to be promoted via Hg-DMSO complexes, and limited when the reactant is a chloro complex (i.e., seawater) due to its inhibitory effect probably because of higher stability 0of the Hg-Cl bond. Therefore, this study emphasized the importance of possible abiotic methylation by a non-dependence upon light in aquatic systems, while the abiotic chemical reactions for methylation are mostly caused by a dependence upon light up to date. In conclusion, this thesis achieved MeHg methylation and demethylation through photochemical reaction in aquatic systems. From the results of this thesis, the site-specific environmental factors i.e. environmental conditions of spatial and temporal variations can be effect on the relative importance of each reaction and the resulting net effect in the aquatic environment. In other words, the reduction of MeHg accumulation possibility in aquatic food chain will be mainly affected by the enhancement of demethylation processes with increasing of UV radiation at the surface waters. Ultimately, the results of this thesis could be a significant contribution to understand the possible photochemical processes and mechanism of Hg demethylation and methylation in water and to estimate the factors that influence the competing methylation and demethylation reactions.Contents Abstract i List of Tables xii List of Figures xiv List of Abbreviations xvi Chapter 1. Introduction 1.1 Backgrounds 1 1.2 Organomercury Compounds 3 1.3 Mercury methylation and demethylation in aquatic environments 7 1.3.1 Mercury methylation processes 9 1.3.2 Mercury demethylation processes 15 1.4 Objectives 24 Reference 28 Chapter 2. Effect of Natural Water Constituents on the Photo-decomposition of Methylmercury and the role of Hydroxyl Radical 2.1 Introduction 38 2.2 Materials and Methods 42 2.2.1 Reagents and sample preparation 42 2.2.2 Photo-reactor and experimental design 43 2.2.3 Analytical methods 46 2.3 Results and Discussion 48 2.3.1 Effect of UV light intensity 48 2.3.2 Effect of pH 52 2.3.3 Effect of Fe3+ ions 55 2.3.4 Effect of NO3- ions 58 2.3.5 Effect of HCO3- ions 62 2.3.6 Effect of DOM 65 2.4 Conclusions 71 References 72 Chapter 3. The Production of Dissolved Gaseous Mercury from MeHg Photo-degradation at Different Salinity 3.1 Introduction 77 3.2 Materials and Methods 80 3.2.1 Sampling and materials 80 3.2.2 Photo-reactor and experimental design 80 3.2.3 Analytical methods 83 3.3 Results and Discussion 85 3.3.1 Effect of UV light wavelength and intensity on MeHg degradation 85 3.3.2 Effect of salinity on MeHg degradation 90 3.3.3 DGM production during MeHg photo-degradation 93 3.3.4 Effect of salinity on DGM production in the presence of nitrate and bicarbonate ions 96 3.3.5 Effect of DOM 101 3.4 Conclusions 104 References 105 Chapter 4. Photochemical Methylation of Inorganic Mercury by Various Organic Compounds 4.1 Introduction 110 4.2 Materials and Methods 114 4.2.1 Materials 114 4.2.2 Photochemical experiments 114 4.2.3 Molecular weight fractionation of DOM experiment 115 4.2.4 Analysis of mercury and other environmental parameters 116 4.3 Results and Discussion 119 4.3.1 Effect of UV irradiation and incubation time 119 4.3.2 Effect of different LMWOMs 123 4.3.3 Effect of pH 125 4.3.4 Effect of DOM derived from litter 128 4.3.5 Effect of DOM-fractions on methylation 132 4.3.6 Reactions between Hg(II) and DMSO in seawater 137 4.4 Conclusion 140 References 142 Chapter 5. Conclusions 5.1 Conclusions 146 5.2 Implications 148 국문초록 151 List of Tables Table 2.1. Photo-decomposition rate constants (RdeMeHg) and half-lives (t1/2) as a function of UVA intensities 51 Table 2.2. Effect of Fe3+ ion concentration on the photo-decomposition rate (RdeMeHg) of MeHg 57 Table 2.3. Effect of NO3- ion concentration on the photo-decomposition rate (RdeMeHg) of MeHg 61 Table 2.4. Effect of HCO3- ion concentration on the photo-degradation of MeHg in the presence of 50 μM Fe3+ ion 64 Table 2.5. Effect of fulvic and humic acid concentrations (mg C L-1) on the photo-decomposition rate (RdeMeHg) of MeHg 67 Table 2.6. Effect of humic acid on the photo-decomposition rate (RdeMeHg) of MeHg in the presence of NO3- ion 70 Table 3.1. The effect of salinity on the rate of MeHg photo-degradation 92 Table 3.2. The effect of salinity on DGM production from MeHg photo-degradation under UVA 98 Table 3.3. The effect of salinity on the rate of MeHg photo-degradation in the presence of nitrate or bicarbonate ions under UVA 99 Table 3.4. The effect of salinity on DGM production in the presence of nitrate or bicarbonate ion under UVA 100 Table 3.5. The effect of DOC concentration on the rate of MeHg photo-degradation and DGM production under UVA 103 Table 4.1. Concentration of MeHg and THg the present of LDOM with and without UVA irradiation 131 Table 4.2. Total organic carbon concentration of the resulting fractions after dialysis 136 Table 4.3. The effect of DOM-fractions on methylation of Hg(II) with UVA irradiation 136 List of Figures Fig. 1.1. Cycling of mercury in aquatic environment 6 Fig. 1.2. Proposed pathway for methylation of mercury in Desulfovibrio desulfuricans 13 Fig. 1.3. Schematic diagram of the overall composition in the dissertation 27 Fig. 2.1. Schematic diagram of experimental design for photo-decomposition of MeHg 45 Fig. 2.2. Effect of UVA intensity on the photo-decomposition of MeHg 49 Fig. 2.3. Effect of pH on the photo-decomposition rate of MeHg 54 Fig. 2.4. Effect of NO3- concentration on the photo-decomposition rate of MeHg 60 Fig. 2.5. Comparison of absorbance spectrum of humic acid and MeHg 68 Fig. 3.1. A schematic diagram of the experimental design used to investigate the photochemical decomposition of MeHg 82 Fig. 3.2. Photo-degradation kinetics of MeHg under UVA and (b) UVB 86 Fig. 3.3. Photo-degradation rate constants as a function of different UV intensities 88 Fig. 3.4. The effect of the MeHg concentration on the rate of photo-degradation under UVA and UVB 89 Fig. 3.5. Dissolved gaseous mercury production from MeHg photo-degradation under UVA and UVB 95 Fig. 4.1. Effect of UV irradiation on the methylation of Hg in the presence of acetate 121 Fig. 4.2. First-order rate plots at different UV irradiation 122 Fig. 4.3. Effect of concentration of methyl donors under UV irradiation on the methylation of Hg 124 Fig. 4.4. Effect of pH on the methylation of Hg in the present of acetate 127 Fig. 4.5. Comparison of fluorescence spectrum of LDOM size-fractionation into three molecular size group 135 Fig. 4.6. MeHg formation of Hg(II) via DMSO in different salinity 139 Fig. 5.1. Possible pathways of MeHg photo-demethylation to enhance and to inhibit in aquatic environments 150 List of Abbreviations CVAFS Cold Vapor Atomic Fluorescence Spectrometer CH3• Methyl radical DGM Dissolved Gaseous Mercury DMS Dimethylsulfide DMSO Dimethylsulfoxide DOC Dissolved Organic Carbon DOM Dissolved Organic Matter EtHg Ethylmercury FA Fulvic acid Fe(OH)2+ Ferrous hydroxide ion Fe3+ Ferric ion HA Humic acid HCO3− Bicarbonate ion Hg Mercury Hg(0) Elemental mercury Hg(II) Divalent mercury Hg22+ Dimeric mercury ion HMWOC High-molecular-weight Organic Compound LMWOC Low-molecular-weight Organic Compound LOI Loss On Ignition MeHg Methylmercury NO2− Nitrile ion NO3− Nitrate ion NOM Natural Organic Matter 1O2 Singlet oxygen OC Organic Carbon OH• Hydroxyl radical OM Organic Matter OOCH3• Peroxomethyl radical r2 Determination coefficient RO2• Organic peroxy radical ROS Reactive Oxygen Species SRB Sulfate-reducing Bacteria SRHA Standard Suwanee River humic acid SRFA Standard Suwanee River Fulvic Acid SRM Standard Reference Matter THg Total mercury US EPA United States Environmental Protection Agency UV Ultra VioletDocto

(A)Study on the degradation and the reduction of acute toxicity of simazine and 4-chloroaniline using photolysis and photocatalysis

Thesis(masters) --서울대학교 대학원 :환경보건학과(환경보건학전공),2008. 8.Maste

The Toxicity Assessment of Explosives Contaminated Soil using Soil Microbial Activity Tests

This study was conducted to determine the toxic effect of TNT and RDX on indigenous soil microbes by measuring enzymatic activity. Denitrification activity, dehydrogenase activity, phosphatase activity, and fluorescein diacetate hydrolytic activity were determined for military firing range, field, and paddy soils exposed to TNT, and RDX from 0 to 1,000 mg/kg and 0 to 4,000 mg/kg, respectively, for 2, 4, and 8 weeks. Soil microbial enzymatic activities decreased with higher TNT and RDX concentration and longer exposure time. Microbial enzymatic activities of firing range soil were higher than field and paddy soils, indicating that indigenous microbes in firing range might have been adapted to TNT and RDX due to pre-exposure of the explosives. In addition, the toxicity of TNT and RDX decreased with higher organic matter because TNT and RDX tend to absorb to soil organic matter. No Observable Effect Concentration (NOEC) values of each microbial enzymatic activity were derived by the geometric mean of NOECs from exposure times (2, 4, and 8 weeks) and soil types (firing range, field, paddy soil). The derived NOECs ranged from 45.3 to 55.2 mg/kg for TNT and 286 to 309 mg/kg for RDX.N

Study on Determination of Permissible Soil Concentrations for Explosives and Heavy Metals

Permissible soil concentrations for explosives (i.e., TNT and RDX) and heavy metals (i.e., Cu, Zn, Pb, and As) heve been derived from human risk and ecotoxicity, respectively. For TNT and RDX, human risk based-permissible soil concentrations were determined as 460 mg-TNT/kg-soil and 260 mg-RDX/kg-soil. Ecotoxicity based-permissible soil concentrations for Cu and Zn were determined from species sensitivity distribution (SSD) and uncertainty factor of 1 to 5, yielding 18.0-40.0 mg-Cu/kg-soil and 46.0-100 mg-Zn/kg-soil. For Pb and As, ecotoxicity data were not enough to establish SSD so that a deterministic method was used, generating 13.8-30.8 mg-Pb/kg-soil and 2.10-4.60 mg-As/kg-soil. It is worth noting that the methodology used to derive permissible concentrations in soil can differ depending on ecotoxicity data availability and socio-economic situations, which results in different permissible concentrations. The permissible concentrations presented in this study have been derived from conservative assumptions for exposure parameters, and thus should be considered as soil standards. In the light of remediation and pollution management of a site of interest, the site-specific and receptor-specific permissible soil concentrations should be derived considering potential receptors, current and future land use, background concentrations, and socio-economic consultation.N

Screening-Level Ecological Risk Assessment for Beneficial Reuse as Soil of Dredged Sediment Contaminated with Heavy Metals

본 연구에서는 중금속의 생태독성자료를 통해 준설토의 육상 재활용을 위한 screening-level 생태위해도 평가를 수행하였다. 대상 중금속 6종(Cu, Zn, Cd, Pb, Cr, Ni)의 독성자료는 USEPA의 ECOTOX를 통해 수집하였으며, 선별과정을 거쳐생태위해도 평가에 사용하였다. 각 중금속의 예측무영향농도는 독성자료의 획득이 가능한 육상 수용체의 종 수에 따라 확률론적 방법(Cu, Zn, Cd)과 결정론적 방법(Pb, Cr, Ni)을 통해 도출하였다. 현장 채취 준설토에서 발견되는 중금속 실제오염농도와 예측무영향농도와의 비교를 통해 생태위해도를 계산하였으며 Cu, Zn, Cr, Pb, Ni의 위해도가 1을 초과하여 생태독성학적위해도가 존재할 수 있는 가능성을 확인하였다. 따라서 재활용 부지의 생태위해도를 고려하였을 때 해당 중금속 오염 준설토를 활용하기 위해서는 중금속 정화 또는 보다 높은 수준의 생태 위해도평가가 선행되어야 할 것으로 판단된다. This study conducted a screening-level ecological risk assessment for heavy metals in dredged sediment for recycling in terrestrial environment. Toxicological information of six heavy metals (i.e., Cu, Zn, Cd, Pb, Cr, and Ni) was collected from ECOTOX of US Environmental Protection Agency, and screened and qualified for the use in the screening-level ecological risk assessment. According to the number of terrestrial ecological receptors for which toxicological information is available, PNEC (Predicted No Effect Concentration) of each heavy metal was derived using either stochastic approach (for Cu, Zn, and Cd), or deterministic approach (for Pb, Cr, and Ni). Hazard quotients of the six heavy metals were derived for a field-collected dredged sediment using the PNEC derived and the PEC (Predicted Environmental Concentration) determined for the dredged sediment. The HQs of Cu, Zn, Cr, Pb and Ni were higher than unity indicating a possibility of ecological risk of the five heavy metals when the dredged sediment is applied in terrestrial environment. Accordingly, remediation processes or a higher-level ecological risk assessment would be needed for the recycling of the material.Y