SOURCES OF DIOXINS TO BALTIC AIR Volatilization and Resuspension As Potential Secondary Sources of Dioxins to Air

Transcription

1 SOURCES OF DIOXINS TO BALTIC AIR Volatilization and Resuspension As Potential Secondary Sources of Dioxins to Air VAN ANH LE Student Degree Thesis in Swedish School of Environmental Chemistry 45 ECTS Master s Level Supervisors: Ian Cousins

2 Volatilization and Resuspension as Potential Secondary Sources of Dioxins to Air VAN ANH LE Supervisor: Ian Cousins Master s Thesis in Swedish School of Environmental Chemistry Department of Applied Environmental Science hoacarla@gmail.com

3 Master s Thesis 211 ABSTRACT Persistent organic pollutants (POPs) are ubiquitous contaminants characterized by semivolatility, low water solubility, high lipophilicity and inherent toxicity. A combination of these properties results in long-rang transport, bioaccumulation and biomagnification through food webs. Elimination of the production, use and emissions of these POPs has been ongoing since the 197s. However, the levels of some POPs are still unacceptably high in some parts of the environment and due to their high persistence levels only decline very slowly over a long period of time. This is especially true for POPs in the Baltic Sea due to long water residence time of approximately 4 years. Numerous studies have been carried out to explore the behavior and fate of the POPs in Baltic regions using analytical methods or modeling approaches. Air-soil exchange plays an important role in controlling the environmental fate of POPs in surface media. Air is a transport medium, which spreads chemicals far away from sources. Soils have received an input of POPs from the atmosphere over a long time period. These chemicals have accumulated in soil solids and, as primary emissions are released, can potentially be rereleased to other environmental media. Therefore, soil could become a significant secondary source of some POPs to the air. In this study, the aim was to determine if volatilization and/or resuspension are potential sources of polychlorinated dibenzo-p-dioxins and dibenzofurans (PCDDs/Fs) ( dioxins ) to Baltic air. Sources of these compounds to Baltic air are particularly interesting because levels of dioxins in fatty fish in the Baltic exceed the levels that are considered fit for human consumption in the European Union guideline. The fugacity quotient approach has been previously shown to be a useful method for exploring the equilibrium status of two connected environmental compartments. Fugacity quotients between the atmosphere and soil are calculated for seventeenth toxic 2, 3, 7, 8,- substituted dioxin congeners. A multimedia mass balance model designed for the Baltic Sea region (POPCYCLING-Baltic) is also employed to study the long-term exchange between air and soil. Estimated fugacity ratios from model simulations are compared with calculated fugacity quotients. Moreover, sensitive analysis is undertaken in order to evaluate the relative effect of background concentration, resuspension and bioturbation transport to the transfer flux from soil to air. I

4 Master s Thesis 211 Fugacities of dioxins in soil are additionally measured directly using equilibrium passive sampling devices. Among available passive samplers, polyoxymethylene 17 µm (POM-17) are chosen to absorb freely dissolved PCDD/Fs molecules in soil. Total soil concentrations are measured to provide input data for the POPCYCLING-Baltic multimedia fate and transport model. Estimated fugacities of dioxins will be compared with directly measured fugacities in soil. The predictive ability of the model is assesses by comparing estimated and measured fugacity. Calculated fugacity quotients showed that lower chlorinated dibenzofuran are close to equilibrium between soil and air while other congeners show disequilibrium. Estimated soil/air fugacity ratios are higher than one but soil still accumulates dioxins because transport process is very slow and non-equilibrium can be maintained for a long period of time. Due to the seasonal variation in concentration, volatilization is higher in summer than in winter. Therefore, net gaseous flux between soil and air can be observed in summer. Sensitivity analysis revealed that volatilization flux is proportional to background soil concentration. High background soil concentration results in high volatilization fluxes and vice versa. The simulation showed that the contribution of resuspension flux to air pollution levels is relatively small in comparison to the influence of variation in background soil concentration. If relatively high and unrealistic resuspension velocities are used as inputs in the model, resuspension is a significant source to the atmosphere. In contrast to background soil concentration and resuspension, bioturbation has no effect on volatilization flux even though high bioturbation rates are used as model inputs. In conclusion, except for light congeners, soil is still a sink of PCDD/Fs present in Baltic air. However, the increase in soil/air fugacity ratios suggest an increasing important of soil-to-air transport in the near future. Equilibrium passive samplers using POM strips are considered as a very simple, reproducible, and inexpensive partitioning method. However, the largest disadvantage of using passive samplers for dioxins is the long time to reach equilibrium. It takes 6 months for PCDD/Fs to obtain equilibrium between soil and POM strips, which exceeded the time for doing a 45 credit thesis. The analytical phase of the experiment is still on-going, and thus it was not possible to include the experimental results in this study. II

5 Master s Thesis 211 Key word: PCDD/Fs, air-soil exchange, volatilization, resuspension, bioturbation, POPCYLING-Baltic model, POM-17 III

6 Master s Thesis 211 ACKNOWLEDGEMENT Having finished my thesis, it is a great pleasure to take an opportunity to all those who accompanied and supported me along the way. First of all, I would like to express my appreciation and gratitude to my extraordinarily supervisor, Assoc. Prof. Ian Cousins for all his support and invaluable advice, in the achievement of my academic goals and my way into scientific world. I am deeply in debt of your endless patience and sympathy that enable me to complete my thesis. It is such luck for me to have you as my supervisor. I would like to address the most special word of thanks to Dr. James Armitage who instructed me from the very early stage of my thesis as well as helped me a lot to stay calm even in the most thrilling moments. You are the brilliant mind-guide who always know how and when to trigger the ideas that pull me out from the state of chaos. I am also grateful for discussions, comments and suggestions from Assoc. Prof. Gerard Cornelissen who provided me with valuable advice on the experimental analysis part. From deep inside, I would like to express my heartfelt thanks to Assoc. Prof. Karin Wiberg for her kindness and helpful during my studies and agreement to be my examiner in this thesis project. To my dear teachers of the Department of Chemistry - Umeå University and Department of Environmental Material - Stockholm University, I would like to express my gratitude to you for all the knowledge and skills I have been taught during this Master s program. As well, I also would like to thank my office-mate Li Zhe for her patience, tolerance and inspiration all the time. From bottom of my heart, it is hard to find a word to express my gratitude to my grandparents, my parents for their care and shares. Family is the most precious treasure that I will give my greatest effort to keep and devote to. The most special thanks to my father who taught me how to pursue my dreams till I achieve them and how to believe in myself. Thank you, Mom, for your big and generous heart that gives me eternal love and caring both in my life and studying. Lastly, I wish to thank Chinh Nguyen, my boyfriend, for his eternal love, encouragement and unyielding support through the process. I also offer my regards and blessings to all my beloved friends who supported me in any respect during the completion of my thesis. Stockholm, May 211 ANH LE IV

7 Master s Thesis 211 TABLES OF CONTENTS ABSTRACT... I ACKNOWLEDGEMENT... IV TABLES OF CONTENTS... V LIST OF FIGURES... VIII LIST OF TABLES... XI ABBREVIATIONS... XII 1. INTRODUCTION... 1 BACKGROUND... 1 THE AIMS OF PROJECT PERSISTENT ORGANIC POLLUTANTS (POPS) DEFINITION, CLASSIFICATION ENVIRONMENTAL FATE CHEMICAL ANALYSIS MODELING DIOXINS Dioxins And Their Physical Chemical Properties Sources And Environmental Fate Degradation Long - Range Transport Bio-Accumulation, Bio-Magnification And Toxicity AIR-SOIL EXCHANGE PROCESSES INVOLVED IN AIR SOIL EXCHANGE Dry Deposition Wet Deposition Volatilization V

8 Master s Thesis 211 Bioturbation Resuspension FACTORS AFFECTING THE AIR-SOIL EXCHANGE PROCESS METHODOLOGY TO STUDY AIR-SOIL EXCHANGE FUGACITY QUOTIENT CONCEPT MULTIMEDIA FATE AND TRANSPORT MODEL OF DIOXINS METHODS TO MEASURE FUGACITY IN SOIL Fugacity Meter Equilibrium Passive Samplers METHODS FUGACITY QUOTIENT POPCYCLING-BALTIC MODEL (VERSION 1.5) Environmental Input Parameters Physical-Chemical Input Parameters Initial Concentrations Alterations To Popcycling/Baltic Model ANALYSIS FUGACITY IN SOIL USING PASSIVE SAMPLER... 3 Sampling... 3 Dry Weight Determination... 3 Development Of Pom-17 Samplers RESULT AND DISCUSSION FUGACITY QUOTIENT CONCEPT MODEL Default Values Sensitivity Analysis EXPERIMENT WITH PASSIVE SAMPLER VI

9 Master s Thesis CONCLUSION RECOMMENDATION APPENDIX A_ALTERATION TO MODEL APPENDIX B_INPUT PARAMETERS APPENDIX C_SIMULATION FOR OTHERS CONGENERS APPENDIX D-SENSITIVE ANALYSIS OF 17 CONGENERS VII

10 Master s Thesis 211 LIST OF FIGURES Figure 1. Important fluxes and partition coefficients (Wiberg et al., 29)... 5 Figure 2. General Structure of PCDDs and PCDFs and numbering of carbon atoms... 8 Figure 3. A schematic of illustration of the sources and environmental fate of PCDD/Fs... 9 Figure 4. A schematic picture of vertical soil aerosol suspension under action of wind (Qureshi et al., 29) Figure 5. The POPCYCLING-Baltic Model aims to quantify the pathways of POPs from the terrestrial environment to the marine environment via the atmosphere and rivers (Wania et al., 2) Figure 6. Compartments in POPCYCLING-Baltic Model (Armitage et al., 29) Figure 7. Illustration of shaking soil with POM Figure 8. Seasonal air fugacity of 2, 3, 7, 8-TCDD Figure 9. Seasonal soil fugacity of 2, 3, 7, 8-TCDD in Swedish Baltic Proper Figure 1. Time trend of air fugacity of 2, 3, 7, 8-TCDD in four Baltic Sea regions Figure 11. Time trend in soil fugacity of 2, 3, 7, 8-TCDD in ten terrestrial regions Figure 12. Fugacity ratios between agricultural soil and air in ten terrestrial regions Figure 13. Seasonal net gaseous fluxes of 2, 3, 7, 8-TCDD in Swedish Baltic Proper Figure 14. Net flux of dioxins in ten terrestrial regions Figure 15. Air Fugacity of 17 Dioxins in Swedish Baltic Proper (A4 west) Figure 16. Soil fugacity of 17 Dioxins in Swedish Baltic Proper Figure 17. Net gaseous fluxes of seventeen congeners in Swedish Baltic Proper Figure 18. Net total flux (µg TEQ h -1 ) of 17 Dioxins in Swedish Baltic Proper Figure 19. Changing in air fugacity of 2, 3, 7, 8-TCDD in Swedish Baltic Proper Figure 2. Changing in soil fugacity of 2, 3, 7, 8-TCDD in Swedish Baltic Proper Figure 21. Net flux of 2, 3, 7, 8-TCDD between air and soil in Swedish Baltic Proper Figure 22. Seasonal net gaseous flux from agricultural soil to the atmosphere in Swedish Baltic Proper 58 Figure 23. Air, soil fugacity and net flux of PECDD in Swedish Baltic Proper Figure 24. Air, soil fugacity and net flux of 1,2,3,4,7,8-HXCDD in Swedish Baltic Proper... 6 Figure 25. Air, soil fugacity and net flux of 1,2,3,6,7,8-HXCDD in Swedish Baltic Proper Figure 26. Air, soil fugacity and net flux of 1,2,3,7,8,9-HXCDD in Swedish Baltic Proper Figure 27. Air, soil fugacity and net flux of HPCDD in Swedish Baltic Proper Figure 28. Air, soil fugacity and net flux of OCDD in Swedish Baltic Proper Figure 29. Air, soil fugacity and net flux of TCDF in Swedish Baltic Proper VIII

11 Master s Thesis 211 Figure 3. Air, soil fugacity and net flux of 1,2,3,7,8-PeCDF in Swedish Baltic Proper Figure 31. Air, soil fugacity and net flux of 2,3,4,7,8-PeCDF in Swedish Baltic Proper Figure 32. Air, soil fugacity and net flux of 1,2,3,4,7,8-HxCDF in Swedish Baltic Proper Figure 33. Air, soil fugacity and net flux of 1,2,3,6,7,8-HXCDF in Swedish Baltic Proper Figure 34. Air, soil fugacity and net flux of 1,2,3,7,8,9-HXCDF in Swedish Baltic Proper... 7 Figure 35. Air, soil fugacity and net flux of 2,3,4,6,7,8-HxCDF in Swedish Baltic Proper Figure 36. Air, soil fugacity and net flux of 1,2,3,4,6,7,8-HpCDF in Swedish Baltic Proper Figure 37. Air, soil fugacity and net flux of 1,2,3,4,7,8,9-HpCDF in Swedish Baltic Proper Figure 38. Air, soil fugacity and net flux of OCDF in Swedish Baltic Proper Figure 39. Compare of soil fugacities, net fluxes of PeCDD (A, B) in different cases Figure 4. Compare of soil fugacities, net fluxes of 1,2,3,4,7,8-HxCDD (A, B) in different cases Figure 41. Compare of soil fugacities, net fluxes of 1,2,3,6,7,8-HxCDD (A, B) in different cases Figure 42. Compare of soil fugacities, net fluxes of 1,2,3,7,8,9-HxCDD (A, B) in different cases Figure 43. Compare of soil fugacities, net fluxes of HpCDD (A, B) in different cases Figure 44. Compare of soil fugacities, net fluxes of OCDD (A, B) in different cases Figure 45. Compare of soil fugacities, net fluxes of 1,2,3,4,7,8-HxCDF (A, B) in different cases Figure 46. Compare of soil fugacities, net fluxes of 1,2,3,7,8-PeCDF (A, B) in different cases... 8 Figure 47. Compare of soil fugacities, net fluxes of 2,3,4,7,8-PeCDF (A, B) in different cases... 8 Figure 48. Compare of soil fugacities, net fluxes of 1,2,3,4,7,8-HxCDF (A, B) in different cases Figure 49. Compare of soil fugacities, net fluxes of 1,2,3,6,7,8-HxCDF (A, B) in different cases Figure 5. Compare of soil fugacities, net fluxes of 1,2,3,7,8,9-HxCDD (A, B) in different cases Figure 51. Compare of soil fugacities, net fluxes of 2,3,4,6,7,8-HxCDF (A, B) in different cases Figure 52. Compare of soil fugacities, net fluxes of 1,2,3,4,6,7,8-HxCDF (A, B) in different cases Figure 53. Compare of soil fugacities, net fluxes of 1,2,3,4,7,8,9-HxCDF (A, B) in different cases IX

12 Master s Thesis 211 Figure 54. Compare of soil fugacities, net fluxes of OCDF (A, B) in different cases X

13 Master s Thesis 211 LIST OF TABLES Table 1. List of POPs under Stockholm Convention (29)... 4 Table 2. Summary of fugacity calculations of different levels of complexity used to describe multimedia contaminant fate (Mackay, 21)... 7 Table 3. TEF schemes for some PCDD/F congeners Table 4. The formulae to calculate fugacity capacity for different compartments (Cousins and Jones, 1998; Mackay, 21) Table 5. Summary of Aspvreten air (Sellström et al., 29) and soil concentrations (Gawlik et al., 2) for selected PCDD/Fs Table 6. Terrestrial and atmospheric compartments in POPCYCLING-Baltic Model Table 7. Half-life of PCDD/Fs in different media (Sinkkonen and Paasivirta, 2) Table 8. Physical chemical properties of PCDD/Fs congeners at 25 C (Aberg et al., 28; Govers and Krop; Trapp and Matthies, 1997) Table 9. Sample preparation Table 1. Henry s law constant at 3 C, organic carbon-water partition coefficient and fugacity capacity in air and soil of 17 congeners Table 11. Calculated fugacity in air, soil and fugacity quotient of 17 congeners Table 12. Sensitivity analysis... 4 Table 13. Formulae to calculate various transport processes within and between air and soil Table 14. Total atmospheric concentration XI

14 Master s Thesis 211 ABBREVIATIONS AOC Amorphous organic carbon BC Black carbon CPW,free Cfree DF DD DOM d.w. EC EMEP H HCB HELCOM HRGC HRMS HxCDD HxCDF HpCDD HpCDF Freely dissolved pore water concentration Freely dissolved water concentration Dibenzofuran Dibenzo-p-dioxin Dissolved organic matter Dry weight European Commission European Monitoring and Evaluation Program Henry s law constant Hexachlorobenzene Helsinki convention High Resolution Gas Chromatography High Resolution Mass Spectrometry Hexachlorinated dibenzo-p-dioxin Hexachlorinated dibenzofuran Heptachlorinated dibenzo-p-dioxin Heptachlorinated dibenzofuran I-TEF Toxic equivalency factors according to NATO/CCMS 1988 I-TEQ Toxic equivalents according to I-TEFs KAW KOA KOW MeOH OC OCDD OCDF OM PAHs PCB(s) Air water partition coefficient Octanol air partition coefficient Octanol water partition coefficient Metanol Organic carbon Octachlorinated dibenzo-p-dioxin Octachlorinated dibenzofuran Organic matter Polycyclic aromatic hydrocarbons Polychlorinated biphenyl(s) XII

15 Master s Thesis 211 PCDD/F(s) PCP PDMS PeCDD PeCDF POC POM POP(s) PUF SPM TCDD TCDF TEF TEQ TOC WHO WHO-TEF WHO-TEQ w.w. μg Polychlorinated dibenzo-p-dioxin(s) and polychlorinated dibenzofuran(s); commonly known as dioxins Pentachlorophenol Polydimethylsiloxane (passive sampler) Pentachlorinated dibenzo-p-dioxin Pentachlorinated dibenzofuran Particulate organic carbon Polyoxymethylene (material used for passive sampling) Persistent organic pollutant(s) Polyurethane foam Settling (or suspended) particulate matter Tetrachlorinated dibenzo-p-dioxin Tetrachlorinated dibenzofuran Toxic equivalency factor Toxic equivalent Total organic carbon World Health Organization Toxic equivalency factor according to WHO; two sets issued, in 1998 and 26 Toxic equivalents according to one of the WHO-TEF sets Wet weight Micrograms (1 μg =.1 mg) XIII

16 Master s Thesis INTRODUCTION Background Industrialization and modernization in recent decades has made a big step in improving our daily life. However, the environment is being threatened with the numerous contaminants released from modern industrial activities. According to the European inventory of existing commercial chemical substances, there are more than 56,72 chemicals used in industry in appreciable quantities. Many of them have been used without thoroughly understanding their physico-chemical properties, fate, and toxicology. An example is the use of persistent organic pollutants (POPs) in the early 2 th century; their detrimental ecotoxicological effects were not realized until the 196s and bans were not introduced until the 197s. POPs are organic chemicals, which are toxic, persistent, bio-accumulative, and susceptible for long-range atmospheric transport (PBT-LRT) (Knut Breiveik, 26). The ability to undergo long-range transport to pristine environments (e.g. Arctic) far away from their emission sources make POPs one of the most problematic environmental issues facing society today. One of regions with high levels of POPs in its ecosystems is the Baltic Sea region, which makes this area one of the most studied sea areas in the world. The Baltic Sea is the largest body of brackish water in the world. The Baltic covers an area of roughly 415 square kilometers. About 16 million people live along the coastline, and a total of 8 million people in the entire catchment area (Helcom, 1993). A large amount of domestic, industrial, and agricultural runoff is discharged into the sea through rivers, outfalls, pipelines, and others effluent points. Harmful and toxic substances, e.g. chlorinated hydrocarbon pesticides (DDT, dieldrin, and endrin), polychlorinated biphenyls (PCBs), polychlorinated dibenzo-p-dioxins (PCDDs), and dibenzofurans (PCDFs) have found their ways into the Baltic Sea. All these substances are toxic to the organisms in the marine environment and probably also to humans due to resistance to degradation and bioaccumulation in marine and terrestrial food chains and webs. The concentrations of PCDD/Fs in fatty fish from the Baltic Sea have exceeded permitted values allowed for human consumption in the European Union (Bignert et al., 27). Therefore, it is important to determine sources of chemicals impacting the sea. A few studies 1

17 Master s Thesis 211 have shown that the bulk of PCDD/Fs accumulated in the Baltic Sea mainly come from atmospheric deposition (Sellström et al., 29; Wiberg et al., 29). Therefore, an understanding of concentration and sources of PCDD/Fs to the atmosphere is necessary in order to build a strategy for risk reduction of dioxins. PCDD/Fs are formed and released in the environment mainly through combustion processes or through the production, use, and disposal of chlorinated aromatic compounds. Accidental fires, volatilization from treated wood factories, recycling plants, contamination of commercial products, etc. are other potential sources to the air. A report for European Monitoring and Evaluation Programme (EMEP) about behavior of PCDD/Fs in air showed that only 1% of the annual PCCD/Fs emissions remain in the atmosphere, about 5% degrade, and 38% are transported outside this region (EMEP, 3/24). The remaining part deposits to other media: about 47%-to soil and vegetation and about 9%-to the sea. Soil has received continuously an amount equivalent to 47% of total annual emission over a period of several decades. Besides, the half-life of PCDD/Fs in soil has been reported to vary from 1 to 15 years, which means that their degradation is very slow under natural conditions. As a result, soil accumulates a significant amount of PCDD/Fs (Cousins and Jones, 1998; Duarte-Davidson et al., 1996; EMEP, 3/24; Harner et al., 1995). It is hypothesized that soil is an important potential secondary source of dioxins to the air in the case of their primary emission reduction (Duarte-Davidson et al., 1996). A study focusing on PCBs has shown that their volatilization from soil is about 5% of the total emission to the atmosphere (Shatalov et al., 21). Lighter PCB congeners have a stronger tendency to move from soil to air than heavier congeners (Backe et al., 24). A study in the UK also claims soil to be a source of PCB and lighter PAHs to the air (Cousins and Jones, 1998). It is therefore hypothesized here that PCDD/Fs present in soils in the Baltic region could potentially be secondary sources to the atmosphere through gaseous transport (i.e. volatilization) or through resuspension of soil solids. To date, we are not aware of any studies conducted in the Baltic region that have examined the potential role of soils as a secondary source of dioxins to the atmosphere. The present study was therefore initiated to explore the central hypothesis using several techniques. 2

18 Master s Thesis 211 The Aims of Project Seventeen (2,3,7,8-substituted) of the 21 congeners of dioxins (21 = 75 PCDDs plus 135 PCDFs) were chosen due to their known high toxicity to mammals and thus potential toxic effects on humans (Kutz et al., 199; Van den Berg et al., 26). Firstly, fugacities are calculated from the physical-chemical properties of dioxins, properties of environmental media and their concentration in each medium. The equilibrium state between soil and air is assessed based on the calculated fugacity quotient. Secondly, a multimedia fate and transport model, used to estimate the fate of POPs in the Baltic Sea region (POPCYCLING-Baltic (Armitage et al., 29)), is applied to obtain estimated fugacities in soil and air as well as longterm fluxes between the two media. Moreover, some sensitive analyses were undertaken in order to investigate the effect of initial soil concentration, bioturbation and resuspension to the rate of transfer from soil to air. Thirdly, fugacities in soil were directly measured using passive sampling devices (polyoxymethylene 17 µm). The aim of this last experiment was to compare measured fugacities with those estimated by the model to assess the model s predictive capability. 2. PERSISTENT ORGANIC POLLUTANTS (POPs) 2.1 Definition, classification Persistent organic pollutants (POPs) are defined as organic substances that are toxic and persistent, could bio-accumulate in food webs, as well as undergo long-range trans-boundary atmospheric transport (Breivik et al., 24; El-Shahawi et al., 21; Lohmann et al., 27). In recent years, attempts have been made to identify the behavior of these substances once released to the environment. Many studies have shown that these chemical do not only bioaccumulate but also bio-magnify in food chains and webs, resulting in adverse health effects to wildlife and humans. The Convention on Long-range Trans-boundary Air Pollutant in 1998 in Aarhus ( Denmark) has provided the basic steps for global and regional control of POPs. In 29 there were 21 compounds which had been listed as POPs by the Stockholm Convention. POPs can be grouped according to their formation and primary origins. POPs can be formed by unwanted by-products of combustion or intentionally produced (Breivik et al., 24; El- Shahawi et al., 21; Lohmann et al., 27). Table 1 summarized the list of present (in 29) POPs as well as their origins and classifications. 3

19 Master s Thesis 211 Table 1. List of POPs under Stockholm Convention (29) Groups Primary Origin POPs Intentionally produced Unintentionally formed as byproducts Pesticides/biocides Industrial chemicals -Specific high temperature environment with presence of chlorines -Combustion derived -Chemical-industrial processes Aldrin, chlordane, chlordecone, dieldrin, endrin, mirex, toxaphene, dichlorodiphenyltrichloroethane(ddt), heptachlor, hexachlorocyclohexane (HCH) including lindane and hexachlorobenzene (HCB) Polychlorinated biphenyls (PCBs) Hexabromobiphenyl (HBBP), perfluorooctane sulfonic acid (PFOS), perflourooctane sulfonyl fluoride(pfos-f), pentachlorobenzene (PeCB) Tetra to heptabromodiphenyl ethers (PBDEs) Polychlorinated dibenzo-p-dioxins and dibenzofurans (PCDD/Fs) Polychlorinated biphenyls (PCBs) Poly-aromatic hydrocarbon (PAHs) Hexachlorobenzene (HCB) Pentachlorobenzene (PeCB) 2.2 Environmental fate The behavior and fate of POPs depends upon their physical chemical properties and the nature of environment they reside in (Wiberg et al., 29). The distribution of POPs in environmental compartments is mainly governed by three equilibrium partitioning coefficients, i.e. the air-water, water-octanol and octanol-air partition coefficients, in which octanol is used as a surrogate for lipid and organic matter (Mackay, 21). POPs are transported between environmental compartments by various transport processes, which are often broadly classified in multimedia models as diffusive and advective transport processes (Figure 1). Diffusive transport between soil and air is a reversible two-way process, comprising dry gaseous deposition, volatilization, sorption and dissolution. Advective transport is the transport of chemical when it present in a moving media, including, in the case of transport between soil an air or vice versa, wet and dry deposition, sedimentation, resuspension, and erosion. Degradation is a pathway to irreversibly remove POPs from an environmental 4

20 Master s Thesis 211 compartment. The most important environmental property controlling soil/air exchange is the organic carbon content of the soil. Due to their high lipophilicity and low water solubility, POPs prefer to accumulate in media with high organic carbon or lipid content. Figure 1. Important fluxes and partition coefficients (Wiberg et al., 29) When present in the atmosphere, POPs can sorb to particles or be present in the gaseous phase due to their semi-volatile nature. POPs are removed from the atmosphere both by physical and chemical processes. Physical removal from the air can occur by wet and dry deposition of vapor and particle-sorbed species. For most organic chemicals, reaction with hydroxyl radical is the dominant degradation process. However, for some compounds, dominant degradation processes could be the reaction with ozone, or nitrate radical or photolysis by sun light. Chemicals associated with particulate matter are suspended not to undergo degradation (Watterson, 1999). Beyond physical chemical removal processes stated above, POPs can undergo biotic degradation in surface water, soil and sediments. Biotic degradation consists mainly of microbial degradation. Abiotic degradation includes hydrolysis, direct and indirect photolysis, and oxidation/reduction reactions. Most POPs accumulate in soil and sediment after deposition from the atmosphere. These accumulated POPs can potentially volatilize back to the atmosphere when levels in the air reduced. On the other hands, POPs in soil can be leached to ground water or be degraded. In water, POPs partition between the particle and 5

21 Master s Thesis 211 dissolved phases. They can also be deposited to bottom sediments or be taken up by aquatic biota. POPs in sediment can be transported back to the water column via diffusion or resuspension in processes analogous to those in soil/air exchange. Another important property of POPs is the potential to undergo long-range transboundary atmospheric transport. POPs can travel a long distance in the atmosphere before depositing on the Earth s surface. Various evidence shows that POPs have been found in remote regions (e.g. Arctic) where they have never been produced or used. Long range atmospheric and ocean water transport are the two main pathways for global transport of POPs, resulting in their ubiquitous presence (Lohmann et al., 27). 2.3 Chemical Analysis The method described here is the method used at Umeå University to analyze chlorinated aromatic compounds in environmental samples. POPs are usually present in very low concentration in background environmental samples. In order to compensate for loss of analyte during extraction and cleanup procedures, isotope labeled recovery standards have been used. Isotope-labeled standards ( 13 C- and 37 Cl-labeled) are added to the samples prior to extraction. Since the analytes and the internal standard in any sample receive the same treatment, the ratio of their signals will be unaffected by any lack of the reproducibility in the procedure. Most of extraction methods for organic pollutants are based on their preference to dissolve in organic solvents. There are various types of extraction techniques and solvents, and the design of the extraction procedure depends on the sample matrix and physical-chemcial properties (e.g. polarity) of the analytes. With gaseous or aqueous samples, solid phase extraction or lipid/lipid extraction is used. For solid samples, Soxhlet extraction or Soxhlet- Dean-Stark extraction is preferred to be used, depending on the water content of samples. After extraction, fat from biological samples and other interfering substances still remain in the samples. Cleanup and fractionation procedures are applied, using dialysis or acid/base columns, or multi-layer columns to separate the analytes from the matrix. The cleaned-up sample extracts then undergo separation and quantification using gas chromatography (GC) combined with mass spectrometry (MS). All the GCs have pressure control of the column and 6

22 Master s Thesis 211 temperature programming of the oven. The MS, which is connected with the GC through an interface, can be low or high resolution. 2.4 Modeling Due to the complexity of understanding chemical fate processes in the environment, as well as the great expense of measuring levels of and conducted experiments on organic pollutants, the interest in developing and applying models for estimating environmental fate is increasing. Many different models have been developed that attempt to describe or predict the fate of chemicals in the environment (Armitage et al., 29; Mackay et al., 1996a; Mackay et al., 1996b; Mackay et al., 1992; Mackay and Wania, 1995; McKone, 1996; Paterson and Mackay, 1989; Sweetman et al., 22; Wania and Mackay, 1995; Wania and Mackay, 1999). The models proposed here for calculating partitioning and behavior of POPs in the environment are based on the standard unit-world fugacity modeling concept as developed by Mackay and co-worker. This multimedia mass balance approach was first developed in the late 197s and it is now widely accepted as a useful, essential tool for understanding of the behavior of POPs in the environment. Table 2. Summary of fugacity calculations of different levels of complexity used to describe multimedia contaminant fate (Mackay, 21) Type of fugacity calculation Level I Key as assumptions -Equilibrium partitioning -Steady state -Closed system Information garnered -General partitioning tendencies for persistent chemicals Level II Level II Level IV -Equilibrium partitioning -Steady state -Opened system -Non-equilibrium partitioning -Steady state -Open system -Non-equilibrium partitioning -Dynamic -Open system -Estimate of overall persistence -Important compartments for removal processes -Relative importance of advection and degradation as removal pathways -Influence of mode of entry on fate and transport -Rates of inter-media transport -Refined assessment of overall persistence and loss pathways -Influence of mode of emission on fate and transport -Time course of respond of contaminant inventory by compartment to any time- varying condition 7

23 Master s Thesis 211 Fugacity models can be used to predict environmental fate of chemicals in a unit world. A unit world is a model world with well-mixed compartments such as air, water, soil, sediment, ect A unit world is supposed to reflect the real world or a part of a real world. Fugacity models increase in complexity from Level I to IV. Level I assumes equilibrium partitioning and is the simplest and possible least realistic type of mass balance while level IV allows timedependent concentrations to be predicted (i.e. it is dynamic model) and may often provide the most realistic type of mass balance. One of the advantages of fugacity models is the ability to increase complexity depending on available information and the requirement of accuracy as well as the purpose of users (Mackay, 21; Mackay and Paterson, 1991; Mackay et al., 1992). A detailed explanation of these different levels is included in table Dioxins Dioxins and their physical chemical properties Dioxins are a group of chlorinated organic chemicals with similar chemical structures. Chlorine atoms can attach to eight different places on two benzene rings, carbon atom 1 to 4 and 6 to 9. The common term dioxins includes 21 congeners, in which 75 congeners are polychlorinated dibenzo-p-dioxins (PCDDs) and 135 are polychlorinated dibenzo-furans (PCDFs). A general chemical structure of PCDDs and PCDFs is presented as Figure 2. Figure 2. General Structure of PCDDs and PCDFs and numbering of carbon atoms Because of the unique environmental properties of dioxins and furans, such as low vapor pressure, extremely low water solubility in water, high lipophilic, resistance to photolytic, biological and chemical degradation, and tendency to bioaccumulation, they are categorized as one of the most harmful organic pollutants. Physical chemical properties of dioxins vary among congeners. In contrast to lipophilicity, vapor pressure and water solubility decrease with increasing the number of chlorine atoms in the corresponding congeners. 8

24 Master s Thesis Sources and environmental fate Dioxins are mainly derived from human activities, but can also be generated naturally by forest fires or volcanic activity. They are not produced for any industrial purpose but unintentionally by-products of numerous industrial and combustion processes. Industrial processes, waste incineration, fuels combustion (wood, coal or oil), chlorine bleaching from pulp and paper mill, and chlorinated pesticides manufacturing were believed to be the main sources of dioxins (Duarte-Davidson et al., 1996). Since the introduction of regulation on dioxins, emission from chlorinated pesticides manufacturing which was historically the biggest source has now become a minor contributor. Therefore, combustion processes have become the most important global contributor to the dioxin source inventory (Deriziotis, 24). In addition, cigarette smoke, home-heating systems, and exhaust from cars also contain small amounts of dioxins. Figure 3. A schematic of illustration of the sources and environmental fate of PCDD/Fs Dioxins enter the environment as mixtures containing a variety of individual components and impurities. Once released to environment, they distribute between environmental compartments as seen in Figure 3. Dioxins can be found in both vapor and particles phases due their semi-volatile nature. Their gas-particle partitioning depends on temperature, amount and nature of particulate matter in the air, and the chlorination of dioxin congeners. The large fraction of the less chlorinated dioxin congeners are present in the gaseous phase in the summer since the temperature is high (Bobet et al., 199; Eitzer and Hites, 1989; Watterson, 1999). 9

25 Master s Thesis 211 Two main pathways by which dioxins are physically removed from the air are wet and dry deposition. When deposited to terrestrial environments, dioxins tend to be associated with soil solids or any surface with a high organic content, such as plant leaves. Large amounts of dioxins accumulate in soil and can be gradually released to other media. Most of the PCDD/Fs deposited from the atmosphere bind strongly to dissolve or particulate organic matter in water. These particles deposit into sediments but can also be transported back to water via resuspension. However, the reverse process is quite slow, resulting in the accumulation of large amounts of dioxins in sediment. This is why sediments are regarded as an important reservoir of dioxins in aquatic environment. Fish and other aquatic biota can uptake PCDD/Fs through diffusion across gills or ingestion of contaminated prey Degradation Photo-degradation can occur to dioxins in the gaseous phase, but mostly not in the particle phase (Brubaker, 1997; Knut Breiveik, 26; Watterson, 1999). Dioxins attached to particulate matter are thought to be resistant to degradation. Less chlorinated compounds are more easily degraded than others (Orth et al., 1989; Pennise and Kamens, 1996). The half-life of PCDD/Fs in the atmosphere was found to be in a wide range from.4 up to 62 hours, depending on light intensity and the chlorination of dioxins. Chemicals in surface waters, which receive much sunlight, have higher rates of removal than bottom water or sediments. The degradation half-live of dioxins in sediments has been estimated to up to 55 days (EPA, Technical Factsheet on Dioxin; Ward et al, 1979), although this may be an overestimate of their degradability. Biodegradation has a minor impact on dioxin destruction because of their high resistance to microbial activity. Volatilization also is not an important removal of dioxins from the water column in comparison to the incorporation in sediments. The most important loss processes for dioxin deposited to terrestrial soils are thought to be photolysis and volatilization. The persistence half-life of TCDD on top soil surfaces may vary from less than one to three years but half-lives in soil interiors may be as long as 12 to 15 years (EMEP, 3/24) Long - range transport Physical and chemical properties of high persistence and semi-volatility, coupled with other unique characteristics of PCDD/Fs, have resulted in their being widely distributed through the 1

26 Master s Thesis 211 global environment, even in remote regions where they have never been used, i.e. Arctic and Antarctic regions. Dioxins can move long distances in the atmosphere before deposition. Dioxins were found in soil and sediments in the Arctic (Brzuzy, 1996; Cleverly D.A. and Carthy, 1996; Oehme, 1993; Wagrowski, 2) Bio-accumulation, Bio-magnification and Toxicity Dioxins are global contaminants due to their toxicity, resistance to degradation, tendency to bio-accumulate and bio-magnify up in the food chain. Dioxins have been detected in mussels, crabs, herring, salmon, guillemot and seal (Kiviranta et al., 23; Rappe et al., 1987; Sakurai et al., 2). Fatty fish caught in the Bothnian Sea (within the Baltic Sea) have exceeded the maximum levels for human consumption in European Union guideline (Bignert et al., 27; Kiviranta et al., 23). Bio-accumulation in such organisms occurs by the ingestion of sediment or by direct uptake of dioxins from water through gill membranes. Since these substances are harmful to aquatic organisms, they threaten the survival of predatory animals and human health. Dioxins can have varying harmful health effects depending on the number and position of the chlorine atoms (Duarte-Davidson et al., 1996; Kutz et al., 199; Van den Berg et al., 26). 2, 3, 7, 8-TCDD or simply TCDD, a molecule with 4 chlorine atoms, is the most toxic dioxin congener. Dioxins are slowly bio-transformed in the body and are not easily eliminated. They tend to accumulate in fat and in the liver. By interacting with a cellular receptor, dioxins can trigger biological effects such as hormonal disturbances and alterations in cell functions. Dioxins and dioxin-like compounds that have the ability to interact with Ah-receptors and cause toxic effects are specified by a toxic factor called Toxic Equivalency Factor (TEF) (Van den Berg et al., 26) as shown in Table 3. This factor indicates the degree of toxicity of each congener compared to 2, 3, 7, 8-TCDD, which is given a reference value of 1. All other congeners are assigned lower TEFs comparable to their relative toxicity. TEF values vary for different species and congeners. The TEF values of individual congeners in combination with their concentration give us the total TCDD Toxic Equivalent (TEQ). To calculate TEQ of a dioxin mixture, the amounts of each toxic compound are multiplied with their TEF values and then summed together. The older International Toxic Equivalent (I-TEQ) and the World Health Organization Toxic Equivalent (WHO-TEQ) are the two available schemes. 11

27 Master s Thesis 211 Table 3. TEF schemes for some PCDD/F congeners (Kutz et al., 199; Van den Berg et al., 26) Congeners WHO-TEF (26) I-TEF (1998) 2,3,7,8-TCDD 1 1 1,2,3,7,8-PeCDD 1.5 1,2,3,4,7,8-HxCDD.1.1 1,2,3,6,7,8-HxCDD.1.1 1,2,3,7,8,9-HxCDD.1.1 1,2,3,4,6,7,8-HpCDD.1.1 OCDD.3.1 2,3,7,8-TCDF.1.1 1,2,3,7,8-PeCDF.3.5 2,3,4,7,8-PeCDF.3.5 1,2,3,4,7,8-HxCDF.1.1 1,2,3,6,7,8-HxCDF.1.1 1,2,3,7,8,9-HxCDF.1.1 2,3,4,6,7,8-HxCDF.1.1 1,2,3,4,6,7,8-HpCDF.1.1 1,2,3,4,7,8,9-HpCDF.1.1 OCDF AIR-SOIL EXCHANGE 3.1. Processes involved in air soil exchange As a result of regulations, the production and use of dioxins as pesticides and herbicides have been prohibited and combustion processes have now become the dominant sources of dioxins in the environment (Cousins and Jones, 1998). Once released into the air, dioxins move away from primary sources before being deposited to terrestrial or water surfaces. Soil can receive inputs of dioxins directly from air deposition or indirectly from plant growing on it. Due to their resistance to biodegradation, the application of herbicide and pesticide containing dioxins the 196s and early 197s still remain in soil today. Soil retains dioxins and thus is considered as a large reservoir of PCDD/Fs, which can potentially be gradually released to the atmosphere or surface waters (Cousins and Jones, 1998). Transport processes between the air and soil play an important role in the accumulation and fate of PCDD/Fs for many reasons. Firstly, one of the main pathways that humans are exposed to dioxins occurs via the agricultural food chain; air-plant-cow-human (Cousins et al., 1999a; Duarte-Davidson et al., 1996). For this reason, the levels of dioxins in air are key in controlling the levels of dioxins in human. Secondly, the atmosphere is the major transport medium for dioxins, controlling the regional and global transport of dioxins. Understanding the exchange 12

28 Master s Thesis 211 processes between air and soil is an important part of studies of the behavior and spreading of dioxins in these environments. Dry deposition The two main processes contributing to air-soil exchange of semi-volatile organic compounds (SVOCs) are : atmospheric deposition and volatilization from the soil (Cousins et al., 1999a). Atmospheric deposition to soil includes dry and wet deposition. If soil is covered with vegetation, it will receive another input from plant decay. Due to its large surface area, vegetation is considered as an effective scavenger of dioxins in the atmosphere present in both particle and gaseous phases (Simonich and Hites, 1995). Dry deposition refers to any physical removal process in the atmosphere that does not involve precipitation (Hemond and Fechner-Levy, 2). There are three dry deposition mechanisms: gravitational settling, impaction and absorption (Hemond and Fechner-Levy, 2). Gravitational settling is a significant removal process for particulate matter with diameter is larger than 1 µm (Kaupp et al., 1994; Mackay, 21). Impaction occurs when air containing particles moves past stationary objects e.g. vegetation or buildings. Some of the airborne particles collide with the objects and stick. Dry deposition of particles depends on the size and density of the aerosol particle, terrestrial surface properties such as roughness and atmospheric conditions such as humidity and wind speed. Atmospheric gases are absorbed by liquid or solid surfaces (soil, vegetation, etc.) (Hemond and Fechner-Levy, 2). The process depends on the physicalchemical properties of the substance, the characteristics of the soil surface (i.e. concentration in soil, roughness and especially the type of vegetation) and the environmental conditions (e.g. wind speed) (Cousins et al., 1999a). Wet deposition Wet deposition refers to processes in which atmospheric chemicals are accumulated in rain, snow, or fog droplets and are subsequently deposited onto Earth s surface. Rain and snow are very efficient scavengers of particles. Compounds are removed from the atmosphere both as vapors (which dissolve in the raindrops) and bound to atmospheric particles (which are incorporate in the rain within or below clouds) (Cousins et al., 1999a). When incorporation of chemicals into water droplets occurs within a cloud (nucleation scavenging), the process is called rainout. When incorporation occurs beneath a cloud (scavenging of particles and gases by droplets), the process is called washout. Gases and vapors in the atmosphere are removed 13

29 Master s Thesis 211 from the air effectively by dissolving into raindrops. Particulate chemicals may also be removed from the atmosphere through wet deposition processes. Particles play a role as nucleation sites from condensation at the onset of water droplet or ice crystal formation. Particles can also be incorporated into already-formed water droplets within a cloud by collision. Removal of particles by rainout is far more effective than dry deposition of particles. The total wet scavenging ratios in the air can be calculated with equation: WT = WP Ф + WG (1-Ф) Where: WT is the total wet scavenging ratios. WP and WG are the sum of the effective scavenging ratios for the substance in the particle and gas phases. Ф is the fraction of chemical in air that sorbed to the aerosol. In conclusion, dry and wet deposition control the deposition of PCDD/Fs to soils. In the case of dry gaseous deposition, PCDD/Fs present in the vapor phase subsequently diffuse into the soil. Association with particles that deposit to soils by gravitational settling or impaction is another pathway. The size of particles is the key parameter determining the dry deposition pathway of PCDD/Fs. However, particle size of PCDD/Fs are not dependent on the degree of chlorination, therefore deposition pathways should be similar for all PCDD/Fs. It is hypothesized that impaction is an important pathway of deposition for PCDD/Fs because enrichment of PCDD/F particles are associated with diameter smaller than.45 µm (Kaupp et al., 1994). In the case of wet deposition, PCDD/Fs are dissolved in precipitation. Alternatively, they are associated with atmospheric aerosols scavenged by precipitation. Deposition is in general dominated by the higher chlorinated congeners, notably octa-chlorinated dibenzo-pdioxins (OCDD), which typically accounts for 2-4% of the total PCDD/F flux (Lohmann and Jones, 1998). Volatilization Volatilization from soil refers to the sum of processes that contribute to the evaporation of a compound from the soil surface and subsequent transport to the atmosphere (Cousins et al., 1999a). In soil dioxins can be sorbed to organic matter (reversibly or irreversibly), leached to ground water, removed by erosion or degraded (biotic or abiotic), or volatilized to the air. With soil covered by vegetation, losses by erosion are less than 1% per year (Mackay, 21). Most of PCDD/Fs remain in the soil at least 9 years because of their high immobility and halflife value (1-15 years) (EMEP, 3/24; Hagenmaier et al., 1992). Predicted soil-water 14

30 Master s Thesis 211 distribution coefficients for dioxins ranging from 1 4 to 1 6 reveal that PCDD/Fs sorb strongly to soils (Brzuzy and Hites, 1995). Net volatilization losses can occur only when the fugacity of the substance in the soil exceeds the fugacity of it in the overlying air. The substance needs to be desorbed from the soil, migrate to the soil surface and then be transferred across the soil/air interface to the air. There are typically three mechanisms to transfer a compound to the soil surfaces. For most SVOCs, the main transport route is through mass transfer with evaporating water. More volatile compounds under very dry conditions, are transferred by upward gas and/or liquid phase diffusion. The main route for compounds that are immobile and highly persistent is soil disturbance (tilling or bioturbation). Bioturbation Many animals spends most of or all their lives below ground seeking food, shelter and mates. Earthworns and other invertebrates usually push their way vertically and horizontally through the soil, displacing particles for short distances away from their bodies. These activity, in turn, yield indirect effects on the volatilization flux of chemicals by transferring chemicals to the surface. A study from McLachlan and co-worker showed that the influence of vertical sorbed phase transport to gaseous exchange between surface soil and the atmosphere is very important for lipophilic compounds (McLachlan et al., 22). Dioxins posses ability to sorb in soil, results in transport via the gas and liquid phases is very slow. Therefore, bioturbation is believed to be another, often neglected, important transport mechanism in the soil. Earthworms bring chemicals to the surfaces by turning over soil layers. When chemicals reach the surface, they need to move across a layer called the stagnant air boundary layer. Substances are transported through this layer by molecular diffusion. The rate of transfer is dependent on the diffusion coefficient and vapor density of substances at the interface. Resuspension Surface soil particles can enter the atmosphere by three different mechanisms, depending on their sizes, as shown in Figure 4. Large particles with diameter > 15 µm can only roll along the surface. That movement of soil particles is called creep. Particles have diameters in the range of 7 to 15 µm have the ability to lift up from the surface. However, these particles are still too heavy to be present in the atmosphere for a long time. Saltation is the phenomenon when particles are suspended from the surface but rapidly fall back. Only particles with diameter smaller than 7 µm can suspend freely under suspension mode. The 15

31 Master s Thesis 211 smaller the particles are, the longer they can remain in the atmosphere. However, very small particles with diameter < 2 µm act like a gas and can only suspend in the air for a few days (Shao, 21). Other factors that affect soil suspension are wind velocity, the roughness of the soil surface and the effectiveness of saltation. Figure 4. A schematic picture of vertical soil aerosol suspension under action of wind (Qureshi et al., 29) 3.2 Factors affecting the air-soil exchange process The soil/air partition coefficient (KSA) is used to describe partitioning of compounds between the air and soil. It can be calculated from these equations (1) Where: foc is the soil organic carbon fraction Kow is the octanol/water partition coefficient H is Henry s law constant Equation (1) shows that the behavior of a compound to partition from soil to air becomes increasingly effective with a higher KOW/H ratio, which indicates the dependence of KSA on the properties of chemical e.g. water solubility, vapor pressure, molecular weight, etc. In addition to physical-chemical properties of a compound, environmental factors also play an important role in controlling the partitioning between air and soil (Cousins et al., 1999a; 16

32 Master s Thesis 211 Duarte-Davidson et al., 1996). These environmental factors are temperature, wind speed, humidity, soil properties, and vegetation cover. SVOCs have a tendency to partition into the particle phase at low temperature. In addition, compounds that exit in the vapor phase are also easily adsorb on solid surfaces at low temperatures thus increasing total deposition of a substance. When the temperature increases 1 C, vapor pressure also increases three to four times. Therefore, higher temperature are usually associated with higher volatilization rates. Increasing wind speed not only proportionally increases the dry gaseous deposition flux but also intensifies volatilization and resuspension (Duarte-Davidson et al., 1996). Relative humidity also has an effect on the volatilization rate. Reducing relative humidity can lead to an increase in the volatilization rate due to loss of water in the soil surface. Soil properties such as organic matter content, moisture, texture, porosity have strong effect on soil-air partition coefficients KSA. According to equation (1), KSA increase proportionally with organic carbon content in soil. Soil moisture content has an effect on volatilization due to increasing the migration rate of a substance to the surface. Low soil-moisture content (.3-.8%) has a strong effect on the soil-air partition coefficient, but KSA is not affected if soil moisture content is from 1.9 to 12% (Hippelein and McLachlan, 2). Soil texture is less important than porosity and moisture content in influencing volatilization losses. The effects of vegetation cover on air-soil exchange are expressed by the Leaf Area Index (LAI) value. LAI is the ratio between the total surface area of the leaves and the groundsurface area a plant or tree occupies. Chemicals can be deposited onto the plant by different processes. They exist on the plant until they are washed off by rain or volatilized. If not removed, they may enter the soil when the plants die and decay into the soil. The deposition flux becomes increasingly effective with higher LAI values. There is a paucity of information in the literature on the effects of vegetation cover on volatilization losses from soil. The vegetation covers and shelters the soil and prevents chemicals from exposure to high temperatures. As a result, the re-volatilization process can be decreased compared to bare soil. However, vegetation can make water evaporate. Chemicals in deeper layers may be transported to the soil surface by convection in the soil water, which may make volatilization losses increase. 17

33 Master s Thesis METHODOLOGY TO STUDY AIR-SOIL EXCHANGE 3.1 Fugacity quotient concept The fugacity concept is used to evaluate the contamination status of environmental media as well as to investigate and predict diffusive transport fluxes (Backe et al., 24; Cousins and Jones, 1998; Duarte-Davidson et al., 1996; Horstmann and McLachlan, 1992). Fugacity can be thought of the fleeing or escaping tendency, and is equivalent to the partial pressure in air (Mackay, 21). A chemical present in two compartments is in equilibrium when their fugacity values in both compartments are equal. The fugacity of a compound in a compartment is calculated from its concentration f = C / (Z*M) Where: C is the concentration of compound in the compartment (g m -3 ) M is the molecular mass (g mol -1 ) Z is the fugacity capacity (mol m -3 Pa -1 ) Table 4. The formulae to calculate fugacity capacity for different compartments (Cousins and Jones, 1998; Mackay, 21) Air Za = 1/RT R is the gas constant (8.314 Pa m 3 mol -1 K -1 ) T is the absolute temperature (K) Fugacity capacity (mol m -3 Pa -1 ) Soil Zs = focρskoczw Koc =.41Kow(*) Water ZW = 1/H (*) according to Karickoff (Mackay, 21) foc is the fraction of organic carbon ρs is the soil density (assumed to be 1.5 g cm -3 for all calculation (Lohmann and Jones, 1998)) Koc is the organic carbon/water partition coefficient Kow is the octanol/water partition coefficient H is Henry s law constant ( Pa m 3 mol -1 ) at 25 C The fugacity capacity, or Z-value, is a measurement of the compartment s capacity to hold or store a given chemical. Its value depends on the properties of media and properties of the chemical. The fugacity capacities of different compartments can be calculated from either thermodynamic theory and/or equilibrium partition coefficients. Table 4 shows the formulae to calculate fugacity capacity of air, soil and water. 18

34 Master s Thesis 211 Temperature correction Since most available physicochemical data are reported at 25 C, it is necessary to adjust parameters which are likely temperature dependent such as KOC and H. KOC is calculated from KOW and these partition coefficients are not usually very temperature-sensitive. However, Henry s law constant is a temperature-sensitive parameter. It is necessary to adjust Henry s law constant value for the sampling site temperature. Temperature correction can be done straightforwardly by using the integrated van t Hoff equation (Backe et al., 24; Beyer et al., 22; Cousins and Jones, 1998). ( ) Where: H1, H2 are Henry s law constants at two temperatures T1, T2 are temperature (K) ΔHaw is the enthalpy of air-water exchange (J mol -1 ) The relative fugacity of two environmental compartments is expressed by fugacity quotients. The fugacity quotients of soil and air are calculated as fs/fa where fs is the fugacity of soil and fa is the fugacity of air. The fugacity quotient concept is a useful method because the fluxes are usually low and difficult to measure experimentally. This concept has been used previously in the literature (Backe et al., 24; Cousins and Jones, 1998; Duarte-Davidson et al., 1996). Fugacity quotient values near one show equilibrium between the two phases. Values which differ from one indicate a tendency for the compound to move from one compartment to the others in attempt to establish equilibrium conditions. When the soil/air fugacity quotient is larger than one, the compound tends to volatilize (i.e. the net gaseous flux is from soil to air) from soil to the air. As the result, the soil may become a secondary source to the air. Beside its convenience, the use of fugacity quotient approach has some limitations. The approach only provides a snapshot for a set of environmental conditions. The partition between air and soil can be affected by many factors such as the distribution of compounds within surface soils, rates of transport and resistance to transport. However, these factors are not accounted for in the calculation of fugacity quotients from field data. Multimedia fate and transport model can help to solve these problems by estimating fugacity quotients as a function of time/temperature etc.. 19

35 Master s Thesis Multimedia fate and transport model of dioxins Several mass balance models have been developed to simulate the exchange of POPs between air and soil (Cousins et al., 1999b; Duarte-Davidson et al., 1996; Harner et al., 1995). For example, a study, which used a two-compartment model, predicted that soil is a significant source of PCDD/Fs to the air (Duarte-Davidson et al., 1996). In UK, each year, the soils released smaller than.15 kg ΣTEQ to the atmosphere. However, it was discussed that the model overestimated soil-air fluxes at that time. They also made a conclusion that the soil would be an important source to the air if the primary sources were reduced in the future. A more complex model applied in Germany showed that background soil and air were in equilibrium. However, for highly polluted soils, desorption from soil was a significant secondary source for atmospheric pollution (Trapp and Matthies, 1997). It should be noted though that it took a long time for dioxin to volatize from soil to air, for instance 2, 3, 7, 8- TCDD was transported.1 m in sandy soil in 12 years (Freeman and Schroy, 1986). Herein, the non-steady state, multi-compartmental, fugacity-based model is employed to simulate the environmental fate of PCDD/Fs in the Baltic Sea. Detailed description of the model is presented in Section Methods to measure fugacity in soil Fugacity meter The exchange of gaseous chemical between the atmosphere and soil is a diffusive process (Hippelein and McLachlan, 1998). The direction and magnitude of the diffusion gradient is determined by the concentrations in the air and soil and by the soil/air equilibrium partition coefficient KSA. KSA can be measured using a solid-phase fugacity meter (Hippelein and McLachlan, 1998; Hippelein and McLachlan, 2). In the fugacity meter, a soil sample is placed inside a glass column through which air is passed. Equilibrium between the air and the surface of soil is established by adjusting the air flow rate passing through the column and comparing measured concentrations in the exhaust air. The output air is collected with a sorbent trap, which is extracted with solvent and analysed on a GC-MS to determine the levels in the exhaust air. The concentration in the soil is also determined. KSA is calculated from the ratio of concentrations in soil and air at equilibrium. The fugacity meter is believed to be a valuable tool for investigating the fate of semi - volatile organochlorine compounds in a solid phases (Horstmann and McLachlan, 1992). Apart from bulky apparatus, this method has some other advantages. KSA is sensitive with temperature, therefore it is necessary to keep 2

36 Master s Thesis 211 temperature stable during the operation of the system. In addition, the rate of the air flow passing through the column need to be adjusted in order to ensure equilibrium between the air and the surface soil Equilibrium passive samplers Freely dissolved concentrations (Cfree) refer to those molecules in an aqueous solution that are not bound to particles or associated with dissolved organic carbon. Cfree can be understood as an effective available concentration for bio-uptake or partitioning. Because it is an effective measure of bioavailability, it is important for assessing the risk associated with a chemical in a compartment. The Cfree of organic contaminants have been successfully measured in the studies of chemical fate and transport (Cornelissen et al., 21; Cornelissen et al., 28a; Jonker and Koelmans, 21). One methodological approach for measuring Cfree that has found widespread use in recent years is the use of equilibrium passive sampling devices. In order to assess the availability of PCDD/Fs in soil, the soil pore-water concentration and total soil concentration of dioxins have been measured. Freely dissolved concentration In this project, polyoxymethylene 17 µm (POM-17) is used to absorb freely dissolved PCDD/Fs molecules in soil. Firstly, soil samples are shaken horizontally with POM-17 for a long time enough to achieve equilibrium. The extraction time for PCDD/Fs achieved equilibrium with soils is still unknown. It is assumed that 6 months is long enough for equilibration of the system. The expected equilibrium time of 6 months used here is based on observations of equilibration time of other SVOCs, e.g. PCBs (1-4 days using POM-17) (Cornelissen et al., 28b), PAHs, PCBs, and PCDD/Fs using POM-55 ( 1-14 days) (Cornelissen et al., 21). In order to check equilibrium status of the system, the POM strips are taken out after 3 and 6 months and stored for analysis. After being sampled, POM strips are cleaned and extracted using liquid chromatographic columns. The samples are injected and quantified on a GC-MS system. The freely dissolved (and available ) soil pore-water concentrations (Cpw, free) are deduced from the chemical contents in the POM samplers (CPOM) with measured passive sampler-water partition coefficients (KPOM) values (Cornelissen et al., 21; Cornelissen et al., 28a; Jonker and Koelmans, 21). 21

37 Master s Thesis 211 Methods used to determined passive sampler-water partition coeffcients outlined by Cornelissen et al. (28a). KPOM are measured by shaking POM with PCDD/F stock solution, without soil. KPOM of 2, 3, 7, 8-substituted congeners are deduced from the KPOM-KOW linear regression of measured KPOM vs. KOW for the non-2, 3, 7, 8-substituted congeners. KPOM for each non-2, 3, 7, 8-substituted congener is measure by shaking POM with stock solution, without soil. A range of methanol-water co-solvent systems is used as substitute for pure water because of the difficulty in measuring PCDD/F concentration at pg to ng per liter in pure water (Cornelissen et al., 28a). In addition, KPOM (non-2, 3, 7, 8-substituted congeners) for pure water is deduced by extrapolating to % methanol. The analytical procedure is similar to the procedure described above. The POM strips are shaken with soil and water until the equilibrium condition is established. Therefore, the fugacity of soil is equal to fugacity of water and fugacity of POM. It can be calculated as follow: (H: Henry s law contants) The calculated fugacity will be compared with the estimated fugacity to assess the predictive ability of the model. POM can accumulate larger amounts of contaminants due to its large surface area. By extracting the plastic phase and concentrating the extract to a small volume, the POM method is able to detect very low aqueous concentration. The method is 4 times more sensitive than a standard 7 µm PDMS-SPME (Jonker and Koelmans, 21). Equilibrium passive samplers using POM strips are considered to be a simple, reproducible, and inexpensive partitioning method. In contrast to active sampling, freely dissolved concentration can be directly measured using POM without dissolve organic carbon (DOC) correction (Cornelissen et al., 21; Cornelissen et al., 29). However, the biggest disadvantage of passive samplers for dioxins is the long time to achieve equilibrium. 22

38 Master s Thesis 211 Total soil concentration Total soil concentrations are also measured so that soil/water partition coefficients (KSW) can be derived. Non-dried soils are extracted with toluene, PUF absorbents and internal standard for 17 hours (Cornelissen et al., 28a; Danielsson et al., 25). The extracts is cleaned using liquid chromatographic columns and analyzed with HRGC/HRMS. Total organic carbon and black carbon contents The freely dissolve aqueous concentration in soil can be affected by the strong binding to organic carbon. The sorption of hydrophobic organic chemicals has been proposed to consist of linear absorption of amorphous organic carbon and nonlinear adsorption of black carbon (Cornelissen and Gustafsson, 24). The concentration in the soil is calculated with CS = faockaoccw + fbckbccw (Cornelissen et al., 28a) KSW = faoc*kaoc + fbc*kbc (1) Where KSW is the soil/water partition coefficient. faoc and fbc are the soil mass fractions of AOC and BC, respectively. KAOC is the AOC-water distribution ratio. KBC is the BC-water distribution ratio. KAOC and KBC is compared to assess the relative important of AOC and BC to sorption. AOCwater distribution ratio is widely calculated using equation: Log KAOC = logkow -(.48 ±.42) (Seth et al., 1999) The soil mass fractions of AOC and BC are analyzed, and thus the BC-water distribution ratio can be deduced from equation (1). The mass fraction of BC can be measure directly as described below while the mass fraction of AOC is the difference between TOC and BC. The methods used for determining TOC and BC are exactly the same procedure presented in Cornelissen et al. (28) and Gustafsson et al. (1997). Total organic carbon is determined with catalytic combustion elemental analysis at 13 C after micro-acidification to remove inorganic carbonates. Black carbon contents were determined by forming a small amount of soil samples into balls and burning samples at 375 C for 18 h in the presence of excess oxygen. 23

39 Master s Thesis METHODS 4.1 Fugacity quotient Air concentrations of PCDD/Fs were taken from a study that was previously undertaken at Aspvreten (south of Stockholm) during the winter of The average temperature at Aspvreten during the winter of was 3 C. The Henry s law constant values were recalculated at this temperature. Soil concentrations were taken from data reported for European reference soils (Gawlik et al., 2). The organic carbon fraction in soil was chosen to be a typical value of.4, in the absence of measured values (analysis was ongoing at the time of writing). The particle and gaseous air concentrations as well as soil concentrations of 17 congeners of PCDD/Fs are shown in Table 5. Table 5. Summary of Aspvreten air (Sellström et al., 29) and soil concentrations (Gawlik et al., 2) for selected PCDD/Fs. Compound Mean air concentration (fg m -3 ) Aspvreten (A2, A3, A4) Pallas Average of (Aspvreten+Pallas) (A1) Mean soil concentration (ng kg -1 ) 2,3,7,8-TCDD ,2,3,7,8-PeCDD ,2,3,4,7,8-HxCDD ,2,3,6,7,8-HxCDD ,2,3,7,8,9-HxCDD ,2,3,4,6,7,8-HpCDD OCDD ,3,7,8-TCDF ,2,3,7,8-PeCDF ,3,4,7,8-PeCDF ,2,3,4,7,8-HxCDF ,2,3,6,7,8-HxCDF ,2,3,7,8,9-HxCDF ,3,4,6,7,8-HxCDF ,2,3,4,6,7,8-HpCDF ,2,3,4,7,8,9-HpCDF OCDF POPCYCLING-Baltic Model (Version 1.5) The non-steady state, multi-compartmental, fugacity-based model employed to simulate the environmental fate of PCDD/Fs in the Baltic Sea used in this study is an adapted version of the POPCYCLING-Baltic model (version 1.5) (Armitage et al., 29). The original 24

40 Master s Thesis 211 POPCYCLING-Baltic model (Wania et al., 2) can be downloaded by following this link The model takes the form of a mass balance statement with expressions for all relevant process rates. The air-soil fugacity ratios provided by the model are compared with the calculated fugacity quotients above. The predictive ability of the model is assessed by comparing model predictions with empirical observations. Figure 5. The POPCYCLING-Baltic Model aims to quantify the pathways of POPs from the terrestrial environment to the marine environment via the atmosphere and rivers (Wania et al., 2). The modification has been previous undertaken by Armitage et al. (29) and enabled the user to define the initial concentration in all compartments. The scenarios for atmospheric concentration can be defined as a function of the initial concentration. Air concentrations are set as the driving function in the model, thus it is not necessary to define any emissions to the model. Enhanced sorption to organic carbon was introduced into the model to account for sorption to black carbon. However, it is currently assumed that there is no enhanced sorption to black carbon. Seasonal variability in atmospheric concentration was taken into account as a sinusoidal function of median value. The model was run in the environment of Visual Basic 6.. Due to essential differences in the properties of toxic PCDD/Fs congeners affecting their environmental behavior, simulations were performed separately for seven 2, 3, 7, 8- substituted dibenzo-p-dioxins and ten 2, 3, 7, 8-substituted dibenzofurans. 25

41 Master s Thesis 211 Environmental Input Parameters Figure 6. Compartments in POPCYCLING-Baltic Model (Armitage et al., 29) The POPCYCLING-Baltic model consists of 85 compartments. Each compartment was considered to be well-mixed (i.e. homogenous) both with respect to environmental and chemical properties. The compartmentalization of the terrestrial (a), marine (b), and atmospheric (c) environment of the Baltic Sea drainage basin in the POPCYCLING-Baltic model are presented in Figure 6. Each terrestrial environment is correlative with its overlying atmospheric compartment, as shown in Table 6. Environmental parameters used are the default parameterizations of the model. Table 6. Terrestrial and atmospheric compartments in POPCYCLING-Baltic Model Terrestrial Region Atmospheric Region Terrestrial Region Atmospheric Region T1 Bothnian Bay A1 North T6 Southern Baltic Coast A3 South T2 Bothnian Sea A1 North T7 Swedish Baltic Coast A4 West T3 Gulf of Finland A2 East T8 Danish Straits A4 West T4 Neva A2 East T9 Kattegat A4 West T5 Gulf of Riga A2 East T1 Physical-chemical Input Parameters Physical chemical properties including phase partition coefficients, the corresponding heats of phase transfer, and first-order rate constants for chemical degradation in different compartment are shown in Table 8. The Henry s law constants, vapor pressures, and water solubilities for the 17 selected congeners were taken from (Govers and Krop). Enthalpy of phase change was taken from 26

42 Master s Thesis 211 (Aberg et al., 28). Three partition coefficients, i.e. octanol-water, air-water, and octanol-air were used to describe environmental phase partitioning. Only two have to provide as input, because the third can be calculated from the other two. Table 7. Half-life of PCDD/Fs in different media (Sinkkonen and Paasivirta, 2) Congeners Half-life times (h) Air Water Soil Sediment 2,3,7,8-TCDD ,2,3,7,8-PeCDD ,2,3,4,7,8-HxCDD ,2,3,6,7,8-HxCDD ,2,3,7,8,9-HxCDD ,2,3,4,6,7,8-HpCDD OCDD ,3,7,8-TCDF ,2,3,7,8-PeCDF ,3,4,7,8-PeCDF ,2,3,4,7,8-HxCDF ,2,3,6,7,8-HxCDF ,2,3,7,8,9-HxCDF ,3,4,6,7,8-HxCDF ,2,3,4,6,7,8-HpCDF ,2,3,4,7,8,9-HpCDF OCDF

43 Master s Thesis 211 Table 8. Physical chemical properties of PCDD/Fs congeners at 25 C (Aberg et al., 28; Govers and Krop; Trapp and Matthies, 1997) Congeners M -logh -log S -logp ΔUAW (g.mol -1 ) (kpa m 3 mol -1 ) (mol l -1 logkow ) (Pa) (Jmol -1 ) 2,3,7,8-TCDD ,2,3,7,8-PeCDD ,2,3,4,7,8-HxCDD ,2,3,6,7,8-HxCDD ,2,3,7,8,9-HxCDD ,2,3,4,6,7,8-HpCDD OCDD ,3,7,8-TCDF ,2,3,7,8-PeCDF ,3,4,7,8-PeCDF ,2,3,4,7,8-HxCDF ,2,3,6,7,8-HxCDF ,2,3,7,8,9-HxCDF ,3,4,6,7,8-HxCDF ,2,3,4,6,7,8-HpCDF ,2,3,4,7,8,9-HpCDF OCDF Where M, H, S, P, K OW and ΔU AW indicate molecular weight, Henry s law constant, solubility in water, vapor pressure, octanol-water partition coefficients and enthalpy of phase change, respectively. 28

44 Master s Thesis 211 Initial concentrations Background soil, sediment concentration were kept as default values used by Armitage et al. (29). The initial sediment concentration is derived from the current sediment concentration (Sundqvist et al., 29) and assuming that it has declined at the same rate as atmospheric concentrations. The EU reference background concentration of soil (Gawlik et al., 2) was employed as the initial soil concentration. There was no distinction between agricultural and forest soil. Initial air concentrations were derived from measurements carried out in Aspvreten (South of Sweden) and Pallas (North of Finland) during the winter of 26/27. The average atmospheric concentrations during winter were obtained from the field measurements. The average atmospheric concentrations in summer were assumed to be lower than the corresponding ones in winter by a factor of 4. Input atmospheric concentration for compartment A2, A3, and A4 were based on measurements undertaken at Aspvreten. The average concentration of Aspvreten and Pallas was applied to compartment A1. The input data for atmospheric concentration are presented in Table 3. The simulation was conducted for a period of 2 years, from 1986 to 26. In that time, atmospheric concentrations were assumed to decrease linearly by a factor of 4 according to measurements taken in pine needles (Armitage et al., 29; Rappolder et al., 27). Alterations to POPCYCLING/Baltic model The previous version of the POPCYCLING-Baltic model adapted by Amitage et al. (29), only focused on the fate of POPs between the atmosphere and marine environment, thus some additional programming was necessary to obtain information about air-soil exchange. Furthermore, some environmental input parameters were changed to investigate the sensitive of volatilization rate, as discussed below. Firstly, initial soil concentrations were changed to evaluate the effect of background soil concentrations. The model was run with the same configuration except for the alteration of initial soil concentration. Background soil concentrations were set at one magnitude lower and one magnitude higher than the default values. Secondly, two advective transport processes that have previously been shown to affect the transfer of chemicals from soil to air were added to the model. Bioturbation helps to physically transport the chemical through the soil layer, which could transport chemicals to the surface where they can volatilize. Resuspension can directly transport chemicals to the 29

45 Master s Thesis 211 atmosphere. Resuspension is the transfer process of chemicals associated with soil particulate matter to the air under the influence of wind. The rates of bioturbation and resuspension are the products of the relevant mass transfer coefficients (MTC), concentrations in the soil, and soil area. MTCs were chosen from literature with default values as 2.3 x 1-8 m h -1 (McLachlan et al., 22) and 6 x 1-1 m h -1 (Qureshi et al., 29) for bioturbation and resuspension, respectively. 4.3 Analysis Fugacity in Soil Using Passive Sampler Sampling Surface soils (-2 cm) were sampled at Aspvreten (south of Stockholm). Agricultural soil was taken from an open field while forest soils were taken in a pine forest nearby. All the samples were kept in dark brown flasks and brought back to laboratory. At the laboratory, soil samples were sieved and homogenized using a sieve with 2 mm diameter. The homogenous soil samples were weighed in cleaned flasks and kept in the freezer to prevent degradation. Dry weight determination Approximately 3. gram aliquots of soil were weighed in small cups for each sample. The small cups were covered with aluminum foil and put in an oven (6 C). After a few days, they were taken out and put in desiccator until they reaching room temperature. The water content of the soil samples was calculated from the difference in weight of the cups before and after drying. Development of POM-17 samplers The passive sampling material (POM) from was pre-cleaned by submerging it in MeOH and then putting it in ultrasonic machine. After one day, POM was taken out of the MeOH and placed in an oven (at 6 C) until they were dry. Sodium chlorine was dissolved with MiliQwater to obtain a solution with 1% (g/g). As shown in Table 9, soil (non-dried, 6 g dry weight) was shaken horizontally in the laboratory with sodium chlorine 1 % (25 ml), POM-17 (.4 g) and NaN3 (.2 g) until they reached equilibrium. POM strips were collected after three months and six months. After equilibration, POM strips were sampled, cleaned with water, and keep in freezer. 3

46 Master s Thesis 211 Table 9. Sample preparation Sample Non-dried soil (g) POM (g) NaN3 (g) VA-GL VA-GL VA-GL VA-GL VA-FS VA-FS VA-FS VA-FS Blk Blk The method used at Umeå University to analyze PCDD/Fs in sediment is described by Sundqvist et al. (29). Sediment, which was spiked with 13 C-labelled internal standards of all 2,3,7,8-substituted PCDD/Fs, was weighed into clean thimbles and extracted with toluene using a Soxhlet-Dean-Stark extractor. The extraction was stopped after 15 hours. The sample was purified with activated copper and fractioned by four open liquid chromatographic columns. The first column contained multiple layers i.e., glass wool, 3g KOH-silica, 3g neutral silica, 6g of 4% (w/w) H2SO4 silica and 3g Na2SO4. N-hexane (6 ml), used to rinse and elute analytes from the column. After evaporation, interfering sulphur present in the extract was removed by adding activated copper. The second column had similar ingredients to the first one, but each absorbent and eluent were only half of the amount. The third column was a glass pipette that was packed with glass wool on either side and a mixture of AX21 carbon (7.9%) and Celite in the middle. The column was first eluted with a mixture of n-hexanedichloromethane (1:4) (4 ml). It was subsequently turned upside down and eluted with 4 ml toluene to collect PCDD/Fs. The elution was transferred to a multilayer silica column containing KOH-silica, silica, 4% H2SO4 silica, and Na2SO4. This last column was eluted with n-hexane. 13 C recovery standards (1,2,3,4-TCDD, 1,2,3,4,6-PeCDF, 1,2,3,4,6,9-HxCDF, and 1,2,3,4,6,8,9-HpCDF) were added to samples before injecting and analyzing on the GC/MS system. Figure 7. Illustration of shaking soil with POM-17 31

47 Master s Thesis RESULT AND DISCUSSION 5.1 Fugacity quotient concept Calculation of fugacity quotient was undertaken for the four atmospheric Baltic regions, namely A1 to A4. A1 regions were separated from the others due to differences in concentrations. Table 1 expresses the results of temperature correction for Henry s law constant and calculation of organic carbon-water partition coefficients, fugacity capacity of air and soil. Fugacity capacity of air is the same for all substances at 25 C. Fugacity capacity of soil is much higher than those of air. Table 1. Henry s law constant at 3 C, organic carbon-water partition coefficient and fugacity capacity in air and soil of 17 congeners. Congeners H KOC ZA ZS 2,3,7,8-TCDD 1.62E E E+8 1,2,3,7,8-PeCDD 1.48E-3 1.3E E+8 1,2,3,4,7,8-HxCDD 1.45E E E+9 1,2,3,6,7,8-HxCDD 1.45E E+7 2.3E+9 1,2,3,7,8,9-HxCDD 1.45E E E+8 1,2,3,4,6,7,8-HpCDD 3.8E+ 1.3E E+6 OCDD 3.29E+ 2.31E E+6 2,3,7,8-TCDF 2.57E+ 1.18E E+4 1,2,3,7,8-PeCDF 2.72E+ 4.1E E-4 1.1E+5 2,3,4,7,8-PeCDF 2.59E+ 5.28E E+5 1,2,3,4,7,8-HxCDF 2.72E+ 1.39E E+5 1,2,3,6,7,8-HxCDF 2.72E+ 1.52E+7 4.2E+5 1,2,3,7,8,9-HxCDF 3.2E+ 2.36E E+5 2,3,4,6,7,8-HxCDF 2.75E+ 1.49E+7 4.6E+5 1,2,3,4,6,7,8-HpCDF 2.85E+ 4.2E+7 1.1E+6 1,2,3,4,7,8,9-HpCDF 3.E+ 6.96E E+6 OCDF 3.11E+ 1.63E E+6 Most of the calculated soil/air fugacity quotients (See table 11) are smaller than one, indicating that the air phase is not in equilibrium with soil phase. This result suggests most of PCDD/Fs have a tendency to remain in the soil. Previous studies also showed a similar result (Cousins and Jones, 1998; Duarte-Davidson et al., 1996). Lighter PCDD/Fs have a stronger 32

48 Master s Thesis 211 tendency to move from soil to air than the heavier congeners. The chemicals with high molecular weight have properties of low ability to volatize and high tendency to partition in organic phase. As expected, 2, 3, 7, 8-TCDF, 1,2,3,7,8-PeCDF, and 2,3,4,7,8-PeCDF has fugacity quotient larger than one, which point out the tendency to volatize from soils. Fugacity quotients of heavier congeners showed that soil is nearly in equilibrium with the atmosphere. Therefore, higher chlorinated congeners still remain in the soil. Table 11. Calculated fugacity in air, soil and fugacity quotient of 17 congeners Congeners f A(2-4) f A1 f S f S /f A(2-4) f S /f A1 2,3,7,8-TCDD 1.4E E E E-3 5.5E-3 1,2,3,7,8-PeCDD 4.4E E E E-4 1.5E-3 1,2,3,4,7,8-HxCDD 4.E E E E-4 5.8E-4 1,2,3,6,7,8-HxCDD 1.2E E E-18 2.E-4 3.4E-4 1,2,3,7,8,9-HxCDD 8.4E-15 5.E E E-3 2.1E-3 1,2,3,4,6,7,8-HpCDD 1.E E E E-1 3.5E-1 OCDD 2.1E E E E-1 2.8E-1 2,3,7,8-TCDF 1.5E E E E+ 7.7E+ 1,2,3,7,8-PeCDF 1.2E-14 7.E E E+ 6.3E+ 2,3,4,7,8-PeCDF 1.9E E E E+ 2.3E+ 1,2,3,4,7,8-HxCDF 1.9E E E E-1 1.6E+ 1,2,3,6,7,8-HxCDF 1.9E E E E-1 1.3E+ 1,2,3,7,8,9-HxCDF 2.4E E E E-1 3.9E-1 2,3,4,6,7,8-HxCDF 2.1E E E E-1 1.6E+ 1,2,3,4,6,7,8-HpCDF 7.E-14 4.E-14 5.E E-1 1.2E+ 1,2,3,4,7,8,9-HpCDF 9.5E E E E-1 4.E-1 OCDF 5.8E E E E-1 3.7E Model To illustrate the use of fugacities and fugacities quotients in the interpretation of the fate of dioxins, we discuss the results for 2, 3, 7, 8-TCDD in the Baltic Proper. The results for all other congeners are listed in Appendix C Default values As mentioned above, seasonality in atmospheric concentration was taken into account in this version of POPCYCLING-Baltic model. Seasonality in air fugacity also follows a sinusoidal function with the highest values in summer and lowest in winter. According to the assumption used in the simulation, atmospheric concentrations decreased linearly by a factor of 4 from 1986 to 26 (Bergknut et al., 21). Figure 8 showed the same trend of air fugacity and changes in concentration. In contrast to the trend in air fugacity, soil fugacity (Figure 9) changed very slowly during the simulation time. As previously discussed, PCDD/Fs are 33

49 Master s Thesis 211 strongly sorbed to soil solids, where the rate of degradation is very slow. In combination with the continuous input from the atmosphere over a very long time period, the level of PCDD/Fs in soil are likely to be stable during the simulation time of 2 years ( ). The spatial distribution of PCDD/Fs is shown clearly in Figure 1 and Figure 11. The fugacity in air of 2, 3, 7, 8-TCDD showed high values in the East, South, and West regions, while it was lower in the Northern Baltic Sea. Figure 1 showed the spartial distribution of 2, 3, 7, 8-TCDD over ten terrestrial regions of Baltic Sea. The abbreviation of T1 to T1 can be found in Table 4. The spatial distribution in air concentration over the Baltic Sea results in the different levels of PCDD/Fs in ten terrestrial regions. 2, 3, 7, 8-TCDD have highest level in T5-Gulf of Riga, lowest in T8-Danish Straits. The model estimated soil/air fugacity quotients in different terrestrial regions are presented in Figure 12. All the fugacity quotients are higher than one, meaning that soil fugacities are higher than air fugacities. Due to disequilibrium between soil and air, dioxins in soil have a tendency to move to the atmosphere until they reach equilibrium. Seasonal net gaseous soilto-air flux shown in Figure 13 revealed that net transfer from agricultural soil to the atmosphere occurs in summer. However, the net flux between soil and air in Figure 14 shows that there was almost no net transfer to the atmosphere. This result can be explained because the exchange between soil and air is not only governed by diffusive transport processes but also the advective transport processes. In this case, the rate of advective transport processes is much larger than the rate of diffusive transport. Therefore, even though soil fugacities are higher than air fugacities, soil is still estimated to be a storage reservoir of dioxins. However, it can also be observed that the fluxes from soil to air increase during the simulation time ( ), and therefore volatilization may become an important processes in the near future. 34

50 Air fugacity 1 16 Pa Soil Fugacity ( 1 16 Pa) Air fugacity ( 1 16 Pa) Master s Thesis Figure 8. Seasonal air fugacity of 2, 3, 7, 8-TCDD Figure 9. Seasonal soil fugacity of 2, 3, 7, 8-TCDD in Swedish Baltic Proper North East South West Figure 1. Time trend of air fugacity of 2, 3, 7, 8-TCDD in four Baltic Sea regions 35

51 Soil Fugacity (1 16 Pa) Master s Thesis T1 T2 T3 T4 T5 T6 T7 T8 T9 T1 Figure 11. Time trend in soil fugacity of 2, 3, 7, 8-TCDD in ten terrestrial regions. f S /f A T1 T2 T3 T4 T5 T6 T7 T8 T9 T1 Figure 12. Fugacity ratios between agricultural soil and air in ten terrestrial regions 6 4 f S /f A Figure 13. Seasonal net gaseous fluxes of 2, 3, 7, 8-TCDD in Swedish Baltic Proper 36

52 Flux (µg TEQ h -1 ) Master s Thesis Figure 14. Net flux of dioxins in ten terrestrial regions T1 T2 T3 T4 T5 T6 T7 T8 T9 T1 The fugacities in air and soil as well as the net gaseous and total fluxes of 17 congeners in the Baltic Proper are shown in Figure 14 to 18. High air and soil fugacities for PCDFs were observed. The occurrence of net gaseous transfer from soil to the atmosphere was observed for the lower chlorinated congeners (e.g. 2,3,7,8-TCDD; PeCDD; TCDF; PeCDF). However, the net total flux from soil to air only occurred to TCDF. Other congeners tend to be close to equilibrium between the air and the soil thus there is negligible net diffusive flux on an annual basic. However, seasonal net gaseous fluxes of these congeners show in Figure 22 reveal a higher volatilization tendency in summer than in winter. A low volatilization flux may occur during the summer period. High lipophilicity means dioxins strongly sorb to organic matter, resulting in their immobility and low degradation. These properties help dioxins accumulate in the soil for a long time period. Besides, the nature of the soil as well as the contamination patterns in that soil also determine the differences in the volatilization flux. 37

53 Soil Fugacity (1 16 Pa) Air FUgacity (1 16 Pa) Master s Thesis TCDD PECDD HXCDD HXCDD HXCDD HPCDD OCDD TCDF PECDF PECDF HXCDF HXCDF HXCDF HXCDF HPCDF HPCDF OCDF Figure 15. Air Fugacity of 17 Dioxins in Swedish Baltic Proper (A4 west) TCDD PECDD HXCDD HXCDD HXCDD HPCDD OCDD TCDF PECDF PECDF HXCDF HXCDF HXCDF HXCDF HPCDF HPCDF OCDF Figure 16. Soil fugacity of 17 Dioxins in Swedish Baltic Proper 38

54 Flux (µg TEQ h -1 ) Master s Thesis TCDD PECDD HXCDD HXCDD HXCDD HPCDD OCDD TCDF PECDF PECDF HXCDF HXCDF HXCDF HXCDF HPCDF HPCDF OCDF Figure 17. Net gaseous fluxes of seventeen congeners in Swedish Baltic Proper TCDD PECDD HXCDD HXCDD HXCDD HPCDD OCDD TCDF PECDF PECDF HXCDF HXCDF HXCDF HXCDF HPCDF HPCDF OCDF Figure 18. Net total flux (µg TEQ h -1 ) of 17 Dioxins in Swedish Baltic Proper 39

55 Master s Thesis Sensitivity Analysis Sensitivity analysis is quantification of changes in model results as a result of changes individual model parameter (McKone and MacLeod, 23). Some sensitivity tests were undertaken to investigate the effect of background soil concentration and advective transport on fugacities in air and soil, rate of transfer from soil to air (Table 12). Table 12. Sensitivity analysis Parameters Case Changing Initial soil concentration A Decrease initial soil concentration ten times B Increase initial soil concentration ten times Bioturbation C Add into model Resuspension D Add into model As a result of defined scenario for the atmospheric compartment, fugacity in air is not affected by changes in the background concentration of soil or by adding bioturbation and resuspension in the model as shown in Figure 19. Figure 2 displays the fugacity of 2, 3, 7, 8- TCDD in soil for all 4 model scenarios listed in Table 12. Firstly, the trend and magnitude of fugacity in soil is not affected by resuspension or bioturbation. Not surprisingly, however, the observed changes fugacity in soil are directly proportional to changes the initial soil concentration. Therefore, if background soil concentration is decreased ten times (Case A), fugacity in soil is also reduced 1 times. The same trend is observed when the initial soil concentration is increased ten times (Case B). Other congeners show a similarly tendency as 2, 3, 7, 8-TCDD, but with various levels due to the differences in background soil concentrations and environmental conditions, e.g. surface covers, temperature. The net flux between soil and air is strongly affected by the assumed background concentration in soil, especially when its concentration is high (Figure 21). With the default value, there is no 2, 3, 7, 8-TCDD transfer to the air, but when the background soil concentration is increased ten times, there is a net flux from soil to air. Soil would become a secondary source of dioxins in this case. Decreasing in soil concentration resulted in a corresponding reduction of volatilization flux. It is observed that the volatilization of lower chlorinated congeners are more sensitive to initial soil concentration than higher chlorinated congeners, which is also reasonable due to their preferences to sorb strongly to soil solids. When resuspension is added to the model, there is a change in the volatilization flux. However, the magnitude of change is smaller than increasing the background concentration 4

56 Master s Thesis 211 and not high enough for soil aerosol resuspension to have a significant influence on volatilization flux. The same result was observed in a study of Qureshi and co-workers (29). When very high mass transfer coefficients is added, the effect of resuspension to soil-to-air transport is significant. High mass transfer coefficient could only be obtained in those regions where are hot, dry, and windy,e.g. desert, regions locates in the lower latitudes. The obtained result is reasonable because the mass transfer coefficient is not considered to be very high in cold climate conditions. The Baltic Sea is located in a cold temperature region just below the Arctic Circle, which has relatively long winters when snow covering the soils prevents soil resuspension occurring. It was shown that the effect of resuspension flux on volatilization varied among congeners. The observed trend revealed that the extent of the effect could depend on the concentration of congeners in the soil, if environmental conditions are the same. Congeners with high soil concentrations showed higher sensitivity to resuspension than others. Similarly, a study of EMEP reported that soil concentrations have influenced on the resuspension flux of PAHs (Gusev et al., 3/28). PCDD/Fs are mostly associated with soil organic matter within the soil compartment. Transport processes associated with soil solids are considered to be more important than diffusive soil-air and soil-water processes. Bioturbation is a often modeled as a diffusive transport of soil solids within soil. In order to volatilize to the air, chemicals must be transferred to the soil/air interface. Chemicals must diffuse through a thin stagnant boundary layer of air above the soil surface. The rate of transfer within the soil combined with the rate of diffusion through this layer are the two key transport processes determined the rate of volatilization. The volatilization is air-side controlled when the rate of diffusion through the stagnant layer is dominant. By contrast, the process is termed to be soil-side controlled if soil transport is dominant. In this study, the sensitivity of bioturbation on soil-to-air transport is examined by changing the mass transfer coefficients for this process. Modeling results shown in Figure 18 to 2 suggest that bioturbation has no effect on the exchange of dioxins between the atmosphere and soil of PCDD/Fs, even though high mass transfer coefficients were added. This suggests that soil-to-air transport is air-side controlled. In others words, diffusion through the atmospheric boundary layer is the main process controlling the air-soil exchange of PCDD/Fs. It has been previously observed that sorbed phase transport has the strongest effect on soil fugacities of chemicals with a log KOA between 7 and 8 and a log KAW > -3 (McLachlan et al., 22). It is clear that the partition coefficients of PCDD/Fs are outside these regions and thus bioturbation has no effect on predicted soil fugacities in the model. 41

57 Flux ( µg TEQ h -1 ) Soil fugacity ( 1 16 Pa) Air fugacity ( 1 16 Pa) Master s Thesis Default values Case A Case B Case C Case D Figure 19. Changing in air fugacity of 2, 3, 7, 8-TCDD in Swedish Baltic Proper Default values Case A Case B Case C Case D Figure 2. Changing in soil fugacity of 2, 3, 7, 8-TCDD in Swedish Baltic Proper Default values Case A Case B Case C Case D Figure 21. Net flux of 2, 3, 7, 8-TCDD between air and soil in Swedish Baltic Proper 42