PRIYANKA CHANDAN MERCURY ISOTOPE FRACTIONATION DURING AQUEOUS PHOTO-

Transcription

1 MERCURY ISOTOPE FRACTIONATION DURING AQUEOUS PHOTO- REDUCTION OF METHYLMERCURY IN PRESENCE OF DIFFERENT TYPES AND AMOUNTS OF DISSOLVED ORGANIC MATTER BY PRIYANKA CHANDAN A thesis submitted in conformity with the requirements for the degree of Master of Applied Science Graduate Department of Geology University of Toronto Copyright by Priyanka Chandan, September 2011

2 MERCURY ISOTOPE FRACTIONATION DURING AQUEOUS PHOTO-REDUCTION OF METHYLMERCURY IN PRESENCE OF DIFFERENT TYPES AND AMOUNTS OF DISSOLVED ORGANIC MATTER Priyanka Chandan Master of Applied Science Graduate Department of Geology University of Toronto, 2011 ABSTRACT The effects of different types and amounts of dissolved organic matter (DOM) on the isotope fractionation of Hg isotopes during aqueous photo-reduction of monomethylmercury (MMHg) were investigated to assess whether mass-independent fractionation (MIF) signatures can be used to track photo-degradation of MMHg in natural waters. From experiments conducted with different amounts of reduced organic sulfur (S red -DOM), it appears that MIF during photo-reduction may be dependent on whether MMHg is dominantly bound to S red -DOM. Similar fractionation factors were observed for experiments where S red -DOM was in far excess of MMHg, while significantly lower fractionation factors were observed with lower S red -DOM. We also characterized the signature of MIF (i.e. Δ 199 Hg/Δ 201 Hg) during MMHg photo-degradation to assess if it was similar in different matrices. The experimental Δ 199 Hg/Δ 201 Hg was very similar for different matrices. However, the experimental slope is slightly but statistically different than the slope observed in freshwater fish, which preserve MMHg in nature. (ii)

3 ACKNOWLEDGMENTS I would like to thank Dr. Bridget Bergquist, my advisor for her academic guidance and support. This thesis would not have been possible without her guidance. I would also like to thank Dr. Barbara Sherwood-Lollar, Dr. Carl Mitchell and Dr. Grant Ferris for their valuable suggestions during my M.A.Sc. Research. I thank Dr. Sanghamitra Ghosh for her assistance, guidance and invaluable moral support during my research. I would also like to thank my family for their love and never-ending support. (iii)

4 TABLE OF CONTENTS Abstract Acknowledgements Table of Contents List of Tables List of Figures ii iii iv vi vii 1. INTRODUCTION Mercury cycle Sources of MMHg in freshwater systems MMHg Methylation Terrestrial Runoff Atmospheric deposition 1.3 Sinks of MMHg in freshwater systems Biological MMHg degradation Abiotic MMHg degradation Sediment burial Outflow 1.4 Dissolved Organic matter Mass-dependent and Mass-independent fractionation Mass-dependent fractionation Mass-independent fractionation MDF and MIF of Hg isotopes in natural and experimental studies Relationship between Δ 199 Hg and Δ 201 Hg MIF of MMHg in fish Quantification of MMHg photo-degradation loss in natural waters Factors affecting MIF of MMHg during photo-degradation in freshwater systems 2. Research Objectives Methods Dissolved organic matter 3.2 Sample Preparation 3.3 Photo-reduction experiments design 3.4 Dark controls 3.5 MMHg concentration analysis 3.6 DOM matrix purification for isotope analysis 3.7 MC-ICP-MS analysis (iv)

5 3.8 Analytical uncertainty 3.9 Rayleigh fractionation 4. RESULTS Concentrations 4.2 Mass-dependent fractionation 4.3 Mass-independent fractionation 4.4 Relationship between Δ 199 Hg and Δ 201 Hg 5. DISCUSSION CONCLUSIONS 63 TABLES 66 FIGURES 72 REFERENCES 89 (v)

6 LIST OF TABLES Table 1. Summary of conditional stability constants for MMHg S red -DOM Complexation 66 Table 2. The elemental composition of organic matter obtained from IHSS 67 Table 3. Summary of six MMHg-DOM experiments and MMHg:S red -DOM ratios 68 Table 4. Rayleigh fractionation factors for MMHg-DOM experiments 69 Table 5. Mercury isotope data for MMHg photo-degradation in presence of different types and amounts of DOM under artificial radiation 70 (vi)

7 LIST OF FIGURES Figure 1. Biogeochemical cycle of mercury 72 Figure 2. MMHg photo-degradation processes in freshwater systems 73 Figure 3. Schematic of the MMHg-DOM photochemical reaction apparatus 74 Figure 4. Schematic of the Gas-liquid phase separator 75 Figure 5: Reaction schematic of MMHg photo-degradation in presence of 76 variety of DOM types and amounts Figure 6. MMHg concentrations in the substrate reservoir over time 77 Figure 7. Mass-dependent fractionation (δ 202 Hg) of the MMHg- SRFA 78 Figure 8. Mass-independent fractionation observed in odd isotopes only during MMHg- SRFA experiments 79 Figure 9. Mass-dependent fractionation (δ 202 Hg) of the MMHg- PL FAR 80 Figure 10. Mass-independent fractionation observed in odd isotopes only during MMHg- PL FAR experiments 81 Figure 11. Mass-dependent fractionation (δ 202 Hg) of the MMHg- NL NOM 82 Figure 12. Mass-independent fractionation observed in odd isotopes only during MMHg- NL NOM experiments 83 Figure 13. Ln (δ 202 Hg) is plotted as a function of ln (f). The slope of the plot can be used to estimate the enrichment factor (ε), which was used to calculate kinetic fractionation factors α δ Figure 14. Ln (Δ 199 Hg) is plotted as a function of ln (f). The slope of the plot was used to estimate the enrichment factor (ε), which was used to calculate kinetic fractionation factors α Δ Figure 15. Comparison of fractionation factors α Δ199 between this study and Bergquist & Blum, (vii)

8 Figure 16. A- Δ 201 Hg versus Δ 199 Hg for photochemical reduction of MMHg in presence of a variety of DOM types and amounts. B- Δ 199 Hg versus Δ 201 Hg is plotted for freshwater fish and Hg (II) photo-reduction and compared to the MMHg-DOM photo-reduction slope in this study. 87 Figure 17: Estimates of MMHg photochemical loss to different choices of fractionation factors i.e. from 20mg/L SRFA experiment from this study under artificial radiation and from 20mg/L SRFA experiment from Bergquist & Blum (2007), carried out under natural sunlight. 88 (viii)

9 1.1 INTRODUCTION Mercury (Hg) is a highly toxic and globally distributed heavy metal that circulates between the atmosphere, terrestrial and aquatic ecosystems (Mason et al., 1994; Fitzgerald and Lamborg, 2004). Hg is emitted into the atmosphere from both natural and anthropogenic sources. Natural sources of Hg include volcanoes, Hg rich ore deposits, geothermal vents, and forest fires (see reviews: Schroeder and Munthe, 1998; Fitzgerald and Lamborg, 2004; Gustin et al., 2008). Anthropogenic sources originate from fossil fuel combustion (i.e. coal, peat, wood), chlor-alkali plants, waste incineration, metal mining and refining and manufacturing processes (see reviews: Schroeder and Munthe, 1998; Fitzgerald and Lamborg, 2004; Xin & Gustin, 2007). Human activities have a huge impact on the increased Hg concentrations in the environment and have resulted in an estimated 3-5 fold increase of Hg concentrations over that of pre-industrial Hg concentrations (Fitzgerald and Lamborg, 2004; see review: Lindberg et al., 2007). In the atmosphere, Hg is released as elemental Hg 0, reactive gaseous Hg (II) (RGM) and particulate Hg (Hg-P) as seen in Figure 1 (Lindqvist & Rodhe, 1985; Schroeder & Munthe, 1998). Among these Hg species, elemental Hg is in the most stable and abundant form (>95%), with a residence time of approximately one year (Schroeder & Munthe, 1998; see review: Fitzgerald & Lamborg et al., 2004). This allows the gaseous Hg to be well mixed in the atmosphere and transported across the Earth s surface (see review: Selin, 2009). Reactive gaseous Hg and particulate Hg, in contrast, have shorter residence times in the atmosphere ranging between days and weeks. These species are more water-soluble than Hg 0 and as such, can be easily removed from the atmosphere via dry and wet deposition (Mason et al., 1994; Schroeder & Munthe 1998; see review: Lindberg et al., 2007). Therefore, 1

10 different Hg species can greatly affect the transport and deposition of Hg in the environment (Lindberg et al., 2007). Atmospheric Hg species are deposited to terrestrial and aquatic systems by dry and wet deposition in both local and global regions (Lindqvist & Rodhe, 1985; Mason et al., 1994). The dry deposition involves direct deposition of Hg 0 without any precipitation to the Earth s surface (Mason et al., 1994; see review: Lindberg et al., 2007). Elemental Hg 0 in the atmosphere can also be removed by oxidation, often photochemical, of Hg 0 to Hg (II) by various oxidants such as ozone, hydroxyl and halogens such as chloride and bromide ions. The RGM and particulate Hg is also rapidly removed from the atmosphere via dry deposition as well as wet deposition such as cloud droplet uptake, rain, and snow (see reviews: Lindberg et al., 2007; Selin, 2009). This is because these Hg species are highly soluble and reactive, and as such, easily scavenged from the atmosphere through wet deposition (Mason et al., 1994; Lindberg et al., 2007). Although the major form of atmospheric Hg deposited to the aquatic systems is divalent Hg 2+ (Mason et al., 1994; see review: Selin, 2009), other species of Hg such as dissolved elemental Hg 0 and organic Hg species also exist in aqueous systems (see review: Ulrich et al., 2001). Hg (II) undergoes a number of biogeochemical processes such as microbial reduction, photochemical reduction (Amyot et al, 1997b; Fitzgerald & Lamborg, 2004) as well as chemical reduction of Hg 2+ species (Allard & Arsenie, 1991). Hg (II) forms strong complexes with organic ligands such that >90% of Hg (II) is bound to dissolved organic matter in freshwaters (see review: Ulrich et al., 2001). In addition, a small fraction of Hg (II) in aqueous systems is methylated by microbial processes (Gilmour & Henry, 1991) to 2

11 produce highly toxic monomethylmercury (MMHg). MMHg comprises of < 10% of total Hg in freshwater, ~ 1 to 1.5% in anoxic bottom water sediments (Boszke et al., 2003) and < 5% in estuarine and marine waters (see review: Ulrich et al, 2001). In the freshwater systems, ~ 72-97% of MMHg is strongly bound to dissolved organic matter (Ulrich et al., 2001). It is a toxic contaminant, which accumulates and bio-magnifies in the aquatic food chains (Schroeder & Munthe, 1998). The concentrations of MMHg in aquatic systems can be controlled by MMHg production processes such as methylation as well as MMHg removal processes such as demethylation, scavenging and outflow (Ulrich et al., 2001; Celo et al., 2006). The main sources of MMHg to freshwater systems consist of natural methylation of inorganic Hg (II) in lake sediments and water column by microbes, terrestrial runoff and atmospheric deposition of MMHg (see reviews: Rudd et al., 1995; Sellers at al., 2001). However, it is also important to recognize the sinks of MMHg in freshwaters systems. Various processes such as abiotic photo-degradation, microbial demethylation, sediment burial and outflow have been suggested as important sinks of MMHg in freshwaters (Sellers et al., 2001). The understanding of the sources and sinks of MMHg in freshwater systems is very important to understanding MMHg in fish and food webs, and are addressed in details in the following sections. 1.2 Sources of MMHg in freshwater ecosystems Methylation of inorganic Hg (II) to MMHg The natural methylation of inorganic Hg 2+ in lake sediments and the water column is considered as an important internal source of MMHg to freshwater systems. The formation of MMHg in anoxic bottom waters, lake sediments and wetlands from inorganic Hg occurs predominantly by biotic processes (Rudd et al., 1995; see review: Ulrich et al., 2001). Jensen 3

12 & Jernelov (1969) were the first to observe the microbial methylation of Hg (II) in lake sediments in Sweden. Sulfate reducing bacteria (SRB s) are considered to be the principal organisms in Hg methylation in freshwater and estuarine anaerobic sediments (see reviews: Ulrich et al., 2001; Celo et al., 2006). Iron-reducing bacteria may also be important for methylation processes in environment (Kerin et al., 2006). The in-lake production of MMHg in the bottom sediments of freshwater systems such as temperate lakes in NW Ontario (Sellers et al., 2001), an arctic Alaskan lake (Hammerschmidt & Fitzgerald 2006) and high latitude lakes of western US (Krabbenhoft et al., 2002) is considered an important source of MMHg to the water column. For example, Sellers et al. (2001) suggested that in-lake MMHg formation in the sediments of Lake 240 of Experimental lake area (ELA) in NW Ontario might contribute as an important internal MMHg source to the freshwater systems. MMHg produced in aquatic systems may depend on the microbial activity, concentration of bioavailable Hg (II) and a number of environmental factors such as temperature, ph, redox potential and presence of inorganic/organic ligands (see review: Ulrich et al., 2001; Hall et al., 2008). Abiotic methylation of Hg (II) can occur but it is considered to be a less important process. It involves the chemical methylation of Hg (II) by various methylating reagents such as methyl iodide and dimethyl sulfide. Trans-methylation reactions result in the transfer of methyl groups from lead and tin alkyls to Hg 2+ for production of MMHg in aquatic systems (Ulrich et al., 2001; Celo et al., 2006). Humic matter (i.e. humic acids and fulvic acids) is also considered an important methylating agent (Ravichandran, 2004). Although the abiotic Hg methylation is possible, its importance and contribution to freshwater systems is yet unknown. 4

13 Terrestrial runoff Terrestrial runoff is considered an important external source of MMHg to freshwater systems (Rudd et al., 1995; Sellers et al., 2001; Grigal, 2002). For example, the MMHg input from terrestrial runoff contributed up to 4% of total Hg input to a drainage lake (Lake 240) in NW Ontario (Sellers et al., 2001). Drainage lakes can receive high concentrations of MMHg from terrestrial runoff (Rudd et al., 1995) Atmospheric deposition Atmospheric deposition can be another external source of MMHg to the aquatic systems (Rudd et al., 1995; Sellers et al., 2001). MMHg in wet deposition was measured to get an estimate of the amount of MMHg deposited in the Great Lakes region in USA. The concentrations of MMHg in rainfall and snow samples ranged between 0.01ng/L to 0.85ng/L (Hall et al., 2005). In another example, ~10% of MMHg input to Lake 240 was estimated to be due to wet deposition (Sellers et al., 2001). Precipitation is considered an important external source of MMHg in seepage and drainage lakes in highly polluted and heavily rainfall regions (Rudd et al., 1995; Ulrich et al., 2001). 1.3 Sinks of MMHg in freshwater systems Abiotic and biotic processes in freshwater ecosystems play an important role in the degradation of MMHg and its removal from aqueous systems (Figure 2) Microbial Demethylation As discussed above, microorganisms can methylate inorganic Hg in anoxic bottom waters, lake sediments and wetlands (Rudd et al., 1995; Ulrich et al., 2001; Selin 2009). However, 5

14 microbes also have the ability to degrade MMHg in the aquatic ecosystems. So far, two important microbial degradation processes have been recognized in the literature. Reductive demethylation of MMHg by Hg resistant microbes containing mer-operons results in the formation of methane (CH 4 ) and volatile Hg 0. These microbes express both mera and merb genes, which code for mercuric reductase and organomercurial lyase enzymes respectively. These genes expressed in Hg resistant bacteria have the ability to reduce inorganic Hg (II) and degrade MMHg. Reductive demethylation involves the removal of methyl group by lyase enzyme to produce methane and Hg (II). Mercuric reductase then, reduces Hg (II) to gaseous elemental Hg 0 that volatilizes to the atmosphere. (Robinson & Tuovinen 1984; Marvin- Dispasquale et al., 2000; Ulrich et al., 2001; Barkay et al., 2003; Kritee et al., 2009). Reductive demethylation predominantly occurs in highly polluted aerobic environments (Barkay et al., 2003; Ulrich et al., 2001). Aside from the mer-mediated MMHg degradation pathway, another bacterial mechanism has been suggested that can demethylate MMHg to methane via dimethylmercury sulfide. This has been observed in sulfate reducing bacteria Desulfovibrio desulfuricans (Baldi et al., 1993; Oremland et al., 1995; Marvin- Dispasquale et al., 2000). Although this is a non-mer demethylation mechanism, both pathways mentioned above are recognized as reductive demethylation (Marvin-Dispasquale et al., 2000). Oxidative demethylation of MMHg by microbes such as sulfate reducing bacteria and methanogens produces carbon dioxide (CO 2 ) and Hg (II) as the end products (Marvin- Dispasquale et al., 2000; Ulrich et al., 2001; Barkay et al., 2003). The biological pathway that results in the oxidative degradation of MMHg is still unknown (Barkay et al., 2003). It has been suggested that Hg (II) formed from oxidative demethylation might act as a substrate 6

15 for re-methylation to MMHg (Barkay et al., 2003), might be complexed with organic ligands or be available to form dimethylmercury (Baldi et al., 1993; Ulrich et al., 2001). Oxidative demethylation has been observed in less polluted anaerobic estuarine and freshwater sediments as well as highly polluted stream and reservoir sediments of Carson River (Oremland et al., 1995; Marvin- Dispasquale et al., 2000; Ulrich et al., 2001) Abiotic MMHg degradation Photochemical degradation of MMHg The photo-degradation of MMHg in freshwater systems is considered an important pathway for the removal of MMHg from surface waters. This process was recognized by Sellers et al., (1996), where they reported that abiotic photo-degradation was responsible for decomposition of MMHg in Lake 240 of ELA in NW Ontario. Since then, a number of studies have been conducted to understand the photodecomposition of MMHg in surface waters and its importance in the MMHg cycling in freshwater systems. Photodecomposition of MMHg has been observed in temperate lake of NW Ontario (Sellers et al., 1996; 2001), high altitude lakes of western US (Krabenhoft et al., 2002), an arctic Alaskan lake (Hammerschmidt & Fitzgerald 2006), boreal lake in NW Ontario (Lehnheer & St. Louis 2009) and recently in the Florida Everglades (Li et al., 2010) and suggests that abiotic photodegradation of MMHg is a significant pathway in aqueous systems. Although the estimates of MMHg photo-degradation loss varies in different natural waters, this process has been reported to account up to ~80% of the MMHg loss from freshwater systems. No loss of MMHg was observed in sterilized lake water and unfiltered water samples in the absence of light, which further supports the abiotic photodecomposition pathway (Sellers at al., 1996; Hammerschmidt & Fitzgerald 2006; Lehnheer & St. Louis 2009; Li et al., 2010). 7

16 Since MMHg loss by photodecomposition is an important sink of MMHg in aqueous systems, various studies have been carried out to estimate the annual loss of MMHg by photodecomposition (Sellers et al., 2001; Hammerschmidt & Fitzgerald 2006; Li et al., 2010). Abiotic, sunlight mediated photo-degradation of MMHg resulted in approximately 78% of MMHg loss in temperate lake (Sellers et al., 2001), about 80% loss of MMHg annually mobilized from Toolik Lake sediments in an arctic Alaskan lake (Hammerschmidt & Fitzgerald 2006) and approximately 32% loss of MMHg input in the Florida Everglades (Li et al., 2010). Until the abiotic photodecomposition of MMHg by Sellers et al., (1996), the estimates of MMHg input to freshwater systems in mass balance studies did not include photo-degradation of MMHg. As such, an important source of MMHg i.e. in-lake MMHg production was considered insignificant and the MMHg inputs to freshwater ecosystems were underestimated (Sellers et al., 1996; Hammerschmidt & Fitzgerald 2006; Li et al., 2010). Various environmental factors such as light intensity, dissolved organic carbon and ph may affect the photodecomposition of MMHg in surface waters. Ultraviolet radiation (UV-A & UV-B) has been suggested to be a significant factor that results in the observed high rates of MMHg photo-degradation in freshwater systems (Krabbenhoft et al., 2002; Lehnheer & St. Louis 2009; Li et al., 2010). The photodecomposition rates of MMHg in oligotrophic lakes such as Lake 240 (Sellers et al., 1996), Lake 239 (Lehnheer & St. Louis 2009) & Toolik Lake (Hammerschmidt & Fitzgerald 2006) and in colored Lake 979 (NW Ontario) increased with an increase in PAR intensity and UV radiation. In clear lakes with low DOC concentrations, photodecomposition of MMHg by UV radiation was limited to 0-1m of the 8

17 surface waters and up to 6m by PAR (Krabbenhoft et al., 2002; Hammerschmidt & Fitzgerald 2006; Lehnheer & St. Louis 2009). This suggests that both PAR and UV radiation are important for MMHg degradation in the surface waters but as lake depth increases, the photodecomposition by PAR becomes relatively more important than UV radiation. In colored lakes, high DOC concentrations limit the UV radiation to few cm of the surface water, which decreases the MMHg photo-reduction rates in aqueous systems. For example, the estimated loss of MMHg calculated from photo-degradation rates of MMHg at high DOC concentrations in the Florida Everglades was determined as 32%, which is considerably lower than the MMHg loss estimates by Sellers et al., (2001) and Hammerschmidt & Fitzgerald (2006). However, if the DOC concentrations were reduced to match the low DOC concentrations of Toolik Lake, the estimated MMHg loss was determined to be approximately 74% of MMHg input in the Florida Everglades (Li et al., 2010), which is similar to estimated loss by Toolik Lake. Chemical degradation of MMHg MMHg can also be chemically degraded by photo-chemically produced reactive oxygen species such as hydroxyl radicals (.OH) (Chen et al., 2003; Hammerschmidt & Fitzgerald, 2010) and singlet oxygen (Zhang & Hsu-kim, 2010). The decomposition of MMHg by hydroxyl radicals is considered an important mechanism in natural waters to remove MMHg from aqueous systems. Chen et al., (2003) showed that hydroxyl radicals produced during nitrate photolysis were responsible for MMHg degradation under laboratory conditions. They also suggested the dissociation of Hg-C bond or Hg-Cl bond as the mechanisms behind chemical decomposition of MMHg. In addition, the photo-reduction of MMHg in Alaskan Toolik Lake has been suggested to occur by hydroxyl radicals produced via photo-fenton 9

18 reactions. However, nitrate photolysis did not play any role in the MMHg reduction in Toolik Lake (Hammerschmidt & Fitzgerald, 2010). Laboratory studies have also shown that degradation of MMHg can also occur by singlet oxygen in presence of natural organic matter (Zhang & Hsu-kim, 2010) Sediment burial The sediment burial fluxes in freshwater systems can be calculated from the mean solidphase MMHg concentration in surface sediments of lakes and net sedimentation rate. It is an important sink of MMHg in aqueous systems. Hammerschmidt et al., (2006) calculated the net sediment burial fluxes of MMHg in Arctic lakes and found the burial fluxes ranged from 0.03 to 0.24 µg/m 2 /y. Sediment burial of MMHg in lake sediments also helps in demethylation of MMHg and as such, act as an important factor that contributes towards MMHg loss in freshwater sediments (Feyte et al., 2011) Outflow Outflow or export of MMHg from aqueous systems is yet another factor that contributes towards MMHg removal from aquatic ecosystems. It is considered an important factor when the distribution and fluxes of MMHg are concerned. Sellers et al., (2001) observed MMHg outflow from Lake 240 of ELA in NW Ontario to be 30% of MMHg external sources. As such, if only the external inputs and outputs were taken into consideration for mass balance, Lake 240 was considered a sink of MMHg. 10

19 1.4 Dissolved Organic matter (DOM) Composition of dissolved organic matter Dissolved organic matter is a heterogeneous mixture of complex organic compounds (Bertilsson & Tranvik, 2000; Ravichandran 2004; Khawaja et al., 2010), which are produced from the decay of bio-matter in all terrestrial and aquatic environments (MacCarthy, 2001). Various factors in the aquatic systems can affect DOM characteristics such as organic matter source, temperature, ph, ionic strength as well as biotic (i.e. microbial) and abiotic (i.e. photochemical) transformation processes (Bertilsson & Tranvik, 2000; Leehneer & Croue, 2003). Freshwater DOM typically consists of ~50% carbon, 35-45% oxygen, 4-5% hydrogen, 1-3% nitrogen and 0.5-1% sulfur (Thurman, 1985; Hintelmann et al., 1997; MacCarthy, 2001). It is made up of 75-80% humic substances and 20-25% of low molecular weight organic acids (i.e. carboxylic acids, amino acids) and carbohydrates (Thurman, 1985; see review: Ravichandran, 2004; Hill et al., 2009) in freshwater systems. Throughout the literature, the humic substances have been characterized into different groups based on various factors. Humic substances can be divided into hydrophobic, hydrophilic and transphilic DOM fractions based on their polarity and into acidic, basic and neutral DOM fractions based on acid/base properties (Leehneer & Croue, 2003; Ravichandran, 2004). Humic substances can also be categorized into humic acids, fulvic acids and humin based on their solubility and adsorption properties (see review: Bansal, 1996) MMHg and DOM complexation Dissolved organic matter is a strong complexing agent, which binds to trace metals and affect their speciation, transformation, transport and bioavailability in the terrestrial and aquatic ecosystems (Xia et al., 1998; see review: Ravichandran, 2004). The major functional 11

20 groups recognized in DOM are carboxylic acids, phenols, amine, alcohols and thiol groups. Carboxylic acids and phenols are the most abundant functional groups in DOM, while thiol groups occur only in trace amounts (Thurman, 1985; Ravichandran, 2004). Thiol ligands are among the reduced sulfur groups present in DOM that play an important role in the complexation of trace metals such as Hg. It has been established through X-ray absorption near edge structure (XANES) and extended X-ray absorption fine structure (EXAFS) spectroscopic studies that Hg (II) preferentially binds to reduced sulfur ligands such as thiol (RSH), disulfide (RSSR), thiol ether (RSR) and disulfane (RSSH) groups in organic matter (Xia et al., 1999 and Hesterberg et al., 2001). Similar spectroscopic studies have been performed for MMHg that provide evidence of MMHg binding to reduced sulfur ligands in dissolved organic matter (Qain et al., 2002; Skyllberg et al., 2003; Yoon et al., 2005). Qian et al., (2002) were the first to present a combined EXAFS and Sulfur K edge XANES spectroscopic study of MMHg in soil and aquatic organic matter and showed that 25-40% of the reduced sulfur ligands were involved in MMHg binding during MMHg-DOM complexation. It has been suggested that MMHg preferentially binds only to thiol ligands in the soil and aquatic organic matter (Amirbahmann et al., 2002; Qian et al., 2002; Skyllberg et al., 2003). This evidence was supported by Yoon et al., (2005), where any binding of MMHg to different complexing ligands such as sulfide/ disulfide and polysulfide complexes was absent or insignificant in soil and aquatic samples. As such, it is believed that MMHg forms 1:1 complexation with thiol ligands, resulting in a linear C-Hg-S co-ordination (Qain et al., 2002; Skyllberg et al., 2003; Yoon et al., 2005). 12

21 Effect of MMHg- S red -DOM ratios MMHg reduced sulfur ligands: EXAFS studies on MMHg binding to thiol groups in the dissolved organic matter have also shown that the MMHg: S red -DOM ratios can greatly affect the binding of MMHg to DOM (Qain et al., 2002; Skyllberg et al., 2003; Yoon et al., 2005; Khwaja et al., 2010). In natural aqueous systems, the concentration of MMHg is very low with respect to the concentrations of reduced sulfur ligands in DOM (see review: Ravichandran, 2004). As such, at low MMHg: S red -DOM ratio, MMHg strongly binds to the reduced sulfur functional groups (i.e. thiol ligands) in DOM. With an increase in the MMHg: S red -DOM ratio, it is believed that a small fraction of MMHg begins to saturate thiol ligands in dissolved organic matter. However, at high MMHg: S red -DOM ratios, when the concentration of MMHg is larger than the concentration of reduced S ligands in DOM, a major portion of MMHg could be complexed to weaker binding sites i.e. O/N functional groups in the DOM since all the reduced sulfur groups are saturated by a fraction of MMHg (Qain et al., 2002; Skyllberg et al., 2003; Yoon et al., 2005). Conditional Stability constants: Haitzer et al., (2002) showed that when Hg is bound to DOM, the measured conditional binding constants values are 23.2 and 10.7 at low Hg: DOM and high Hg: DOM ratios respectively. There are various studies that determine the conditional stability constants at low MMHg: S red -DOM ratios in both soil and aquatic systems. It has been reported that at low MMHg: S red -DOM ratios, where MMHg preferentially binds to the reduced sulfur ligands, the conditional stability constants i.e. log K values have been reported to range from 10.7 to 17.1 (Hintelmann et al., 1995; 1997; Amirbahmann et al., 2002; Karlsson & 13

22 Skyllberg et al., 2003; Khwaja et al., 2010; Dong et al., 2010). These stability constants, however, are much lower than the stability constants for Hg (II) DOM complexation (Haizter et al., 2002; Amirbahmann et al., 2002; Dong et al., 2010). However, for high MMHg: S red -DOM ratios, log K values of (Libich & Rabenstein 1973) and (Rabenstein et al., 1974) have been reported for MMHg carboxyl ligands and MMHg amine ligands respectively. Table 1 provides a summary of conditional stability constants for the complexation of MMHg to reduced sulfur, carboxyl and amine ligands in the dissolved organic matter (Dong et al., 2010). In addition to the concentration of MMHg and reduced sulfur ligands (i.e. MMHg: S red ratios), the conditional stability constants can also be affected by ph. At low ph, binding strength between MMHg and DOM is weaker (Khwaja et al., 2010; Karlsson & Skyllberg et al., 2003) and as such, the concentration of free MMHg in soil and aquatic OM increases with a decrease in ph. This might result in the release of MMHg to aqueous systems (i.e. acidic lakes) where it would be available for various processes (Hintelmann et al., 1995). 1.5 Stable isotope fractionation: Stable isotope fractionation has been used as an important tool to identify various sinks, sources and pathways in biogeochemical cycles. The possibility to understand and trace the various transformation processes in the biogeochemical cycle of heavy elements such as Mg, Ca, Fe, Cr, Cu, Zn and Hg has only been possible due to the development of high precision, new analytical techniques such as the multi-collector inductively coupled plasma mass spectrometer (MC-ICP-MS). The observation of stable isotope fractionation of Hg by using MC-ICP-MS was first published in 2001 (Lauretta et al., 2001) and this new field is rapidly growing. 14

23 Stable isotope fractionation happens due to numerous processes. In most equilibrium and kinetic reactions, isotopic fractionation is driven by the relative differences in atomic mass and the effects these can have on vibrational and kinetic energies (Bigeleisen & Mayer, 1947; Urey, 1947). Although kinetic and equilibrium reactions follow slightly different laws, the fractionations can be predicted using mass and thus are called mass-dependent fractionation (MDF). MDF was predicted and has been observed for more than 60 years. Although less common, isotopes can also fractionate in such a way that is not necessarily dependent on mass. Thus, these fractionations are termed mass-independent fractionation (MIF). MIF was first observed for oxygen in the mid 1980s (Thiemens and Heidenreich, 1983). MIF can be caused by numerous processes including molecular symmetry, selfshielding, nuclear volume, nuclear spin, and magnetic moments (see reviews: Thiemens, 2006). For Hg, it is thought that the two major causes of MIF are the nuclear volume effect (Biegeleisen, 1996) and the magnetic isotope effect (Buchachenko, 2001) Mass-dependent isotope fractionation: Mass-dependent isotope fractionation occurs in both equilibrium processes, such as isotope exchange reactions, and in kinetic processes, which depend on the differences in the rates of reaction of different isotopes (see reviews: White, 2005 & Hoeffs, 2006). It has been established that mass-dependent isotope effects mostly depend on the atomic mass of the isotopes and the effect of mass on vibrational frequencies (see review: Faure & Mensing; White, 2005; Hoeffs, 2006; Criss & Farquhar 2008). As such, differences in isotope masses of the same element can result in a number of physical and chemical differences in density, temperature, vapor pressure, melting and boiling points (see review: Hoeffs, 2006). These 15

24 differences can be calculated through quantum mechanic theory (Bigeleisen & Mayer, 1947; Urey, 1947; see review: Young et al., 2002). The differences in the vibrational energy of bonds in molecules with different isotopes are considered to be cause of equilibrium isotope fractionation (see reviews: Clark & Fritz, 1998; Hoeffs 2006). According to quantum mechanical theory, molecules at ground state and absolute temperature have zero point vibrational energy (ZPE), which is ½hν above the minimum potential energy curve where h is the Plancks constant and v is the vibrational frequency of the molecule (White 2005; Hoeffs 2006). Different isotopic species of an element have different ZPE and vibrational frequencies. The ZPE and frequencies depend on mass of the isotopes such that heavier isotopes have lower zero point vibrational energies and vibrational frequencies while lighter isotopes have higher zero point vibrational energies and frequencies. Hence, different masses of isotopes affect the vibrational energy and frequency, which in turn affects the chemical bond strength of the molecules with different isotopes of an element. As such, heavier isotopes favor molecules with strong chemical bonds because the substitution of heavier isotopes in molecules increases the strength of the chemical bonds, lowers the vibrational energies and makes lowers the overall energy of the system. Lighter isotopes, on the other hand, preferentially will be found in molecules with weak chemical bonds (see reviews: Clark & Fritz, 1998; White 2005; Hoeffs, 2006; Criss & Farquhar, 2008). Kinetic isotope fractionation occurs in fast, incomplete or unidirectional processes such as evaporation, dissociation reactions, diffusion and biotic reactions. As mentioned above, heavier isotopes have lower zero point vibrational energies and frequencies. As such, the 16

25 dissociation energy needed to break the bonds and overcome the activation energy barrier is higher for heavier isotopes and lower for lighter isotopes (see reviews; White, 2005; Hoeffs, 2006). This difference in the energy to dissociate the molecules containing heavier and lighter isotopes results in different reactions rates (Clark & Fritz, 1998). Molecules containing lighter isotopes dissociate faster than the molecules with heavier isotopes and the velocity of the lighter isotopes is faster than the heavier isotopes. As such, the heavier isotopes are preferentially retained in the reactants while the products are enriched with lighter isotopes (see reviews: White 2005; Faure & Mensing, 2005). The mass-dependent isotopic compositions of Hg are reported in delta notation with per mil ( ) deviations from NIST 3133 standard as shown in Eq (1). δ X Hg = [( X Hg/ 198 Hg) sample /( X Hg/ 198 Hg) NIST3133 Std ) 1] x 1000 (1) In stable isotope geochemistry, lowest mass isotope is used in the denominator to report isotope ratios. For Hg isotopes, 202 Hg/ 198 Hg ratio (i.e. δ 202 Hg) has been suggested to report MDF Mass-independent isotope fractionation: Mass-independent fractionation of Hg isotopes has been suggested to occur by the nuclearvolume effect (NVE) and the magnetic isotope effect (MIE). These isotope effects are dependent on nuclear properties such as nuclear spin, size and shape, which result in fractionation of only the odd Hg isotopes that deviates from the mass-dependent fractionation. 17

26 Magnetic isotope effect: The magnetic isotope effect (MIE) is a kinetic isotope phenomenon, which results from the interactions between the non-zero magnetic moment of nuclei and magnetic moment of unpaired electrons (see reviews: Turro 1983; Buchachenko, 2001). These interactions are known to occur in chemical reactions containing radical pairs (see reviews: Buchachenko 2001; Buchachenko et al., 2007). Although these interactions are very small, they result in a kinetic isotope effect because of changes in the rate of chemical reactions due to different nuclear spins of isotopes, resulting in the fractionation of magnetic and non-magnetic isotopes into different pools either the reactants or products (see reviews: Turro 1983; Buchachenko et al., 2001, 2007). In chemical reactions, MIE can be demonstrated by the radical pair mechanism. Processes such as photo-dissociation can produce excited molecules in triplet states known as triplet radical pairs (Buchachenko et al., 2007) as shown in Figure 5. Once a triplet radical pair is formed, the recombination of the radical pair into a singlet state molecule with an electronic spin of zero is spin-forbidden (see reviews: Turro 1983; Buchachenko, 2001; Buchachenko et al., 2007). The presence of odd isotopes in triplet radical pair results in the interactions between odd isotope nuclei and unpaired electrons. As such, intersystem crossing (ISC) between non-reactive triplet state to reactive singlet state can occur where odd isotopes undergo a spin conversion. This ISC for odd isotopes is relatively faster than the even isotopes. As such, the singlet radical pair preferentially recombines and results in the accumulation of odd isotopes in the reactant. In the case of even isotopes, the triplet radical pair undergoes a slower intersystem crossing such that the radical pair preferentially dissociates into free radicals and the products are depleted in odd isotopes. Thus, there is a 18

27 difference in the rate of ISC between the odd and even isotopes (see reviews: Turro 1983; Buchachanko 2001; Bergquist & Blum 2009; Buchachenko et al., 2007; Buchachenko 2009). Mass-independent isotopic compositions of Hg are reported with capital delta notation, which is the deviation in the isotope ratios from the theoretically predicted isotope ratios by MDF. MIF is reported in units of per mil ( ). Δ X Hg = δ X Hg - (β x x δ 202 Hg) (2) The β values used are for kinetic mass dependent fractionation of the Hg isotopes (Young et al., 2002) The MIE was first observed by Buchachenko et al. (1976) for carbon isotopes in the photolysis of dibenzylketone (DBK). Since then, there have been numerous studies that have observed MIE in oxygen, sulfur, silicon, germanium and uranium (see review: Buchachenko 2001). Recently, MIF was observed in the abiotic photochemical reduction of Hg and MMHg in aqueous systems by Bergquist & Blum, 2007, who suggested that MIF observed in Hg isotopes may be due to magnetic isotope effect. Nuclear Volume isotope effect Nuclear volume effect is another isotope effect that may result in mass-independent fractionation of odd isotopes. It is due to differences in nuclear volume and nuclear charge radius that result in different ground electronic energies of isotopes of an element. NVE has been theoretically studied for heavier isotopes such as U (Bigeleison, 1996), Tl and Hg (Schauble 2007). For large nuclei, the nuclear volume and nuclear radius is non-linearly 19

28 dependent on the number of neutrons and mass. It has been shown that odd isotopes of Hg such as 199 Hg and 201 Hg have smaller nuclear charge radii than would be predicted by mass. As such, this results in odd-even staggering of isotopes where the odd mass isotopes have nuclear volumes and charge radii closer to the adjacent lower mass even isotope (Bigeleison, 1996; Schauble 2007). Differences in nuclear charge radii can affect the bond strengths and therefore result in isotope fractionation. MIF due to the NVE for 196 Hg, 199 Hg, 201 Hg and 204 Hg was first theoretically predicted by Schauble (2007), but new estimates have been made by Ghosh et al. (2008), Estrade et al. (2009) and Wiederhold et al. (2010) using different compilations of nuclear charge radii that only predict MIF for 199 Hg and 201 Hg Mass-dependent and mass-independent fractionation of Hg isotopes Hg is a highly toxic pollutant in the environment. It has an atomic mass of amu and an electronic configuration [Xe] 4f 14 5d 10 6s 2. There are seven stable isotopes of Hg i.e. 196 Hg, 198 Hg, 199 Hg, 200 Hg, 201 Hg, 202 Hg and 204 Hg, with a relative mass difference of 4%. 202 Hg has the highest natural abundance of 29.86% and 196 Hg has the lowest natural abundance of 0.15%. It has various properties that can lead to isotope fractionation such as active oxidation-reduction chemistry with three oxidation states, ability to form covalent bonds, existence of volatile Hg species (Hg 0 ) and biological cycling of Hg. The even Hg isotopes have a zero spin (i.e. angular momentum) and no magnetic moment. As such, these are called non-magnetic isotopes. The odd isotopes of Hg, on the other hand, have a non-zero spin and magnetic moments (i.e magnetic isotopes). Hg 199 has a spin of -1/2 and magnetic moment of while Hg 201 has a spin of -3/2 and magnetic moment of (Buchachenko, 2001). 20

29 In order to interpret Hg isotope signatures in nature, it is necessary to constrain and quantify the processes that fractionate Hg isotopes in nature. For various natural samples, Hg isotopes display only MDF including source rocks (Smith et al., 2008), hydrothermal systems and ore deposits (Smith et al., 2005, 2008; Sherman et al., 2009), chondrites (lauretta et al., 2001) and volcanic emissions (Zambardi et al., 2009). In addition, there are numerous experimental studies that have explored Hg isotope fractionation during abiotic transformations of Hg such as volatilization of Hg 0 (aq) to Hg 0 (g) (Zheng et al., 2007) and experimental biotic reduction of Hg (II) and MMHg (Kritee et al., 2007, 2008, 2009) that only display MDF. However, for a small subset of samples, Hg also displays MIF. The MIF for Hg isotopes was first observed during aqueous photochemical reduction of Hg (II) and MMHg in presence of dissolved organic matter (Bergquist & Blum, 2007). Both MDF and MIF has been observed for natural samples such as fish and food web samples (Bergquist & Blum, 2007, 2009; Jackson et al., 2008; Gantner et al., 2009; Laffont et al., 2009; Gerhke et al., 2010), coal (Biswas et al., 2008), mosses (Ghosh et al., 2008), lichens (Carignan et al., 2009; Bergquist & Blum, 2009), peat (Ghosh et al., 2008; Biswas et al., 2008), sediments and soils (Biswas et al., 2008; Gehrke et al., 2008; Gantner et al., 2009) and Arctic snow (Sherman et al., 2010). Experimentally, large MIF (>0.5 ) appears to only be associated with photochemical reactions of Hg. Under laboratory conditions, large MIF signatures have been reported during abiotic photochemical transformation processes of Hg such as photochemical reduction of Hg in aqueous systems (Bergquist & Blum, 2007; Zheng & Hintelmann, 2009; Zheng & Hintelmann, 2010a) and snow (Sherman et al., 2010), photochemical reduction of MMHg in aqueous systems (Bergquist & Blum, 2007) and acidic & saline aqueous solutions (Malinovsky et al., 2010). However, smaller MIF are also observed in a variety of other 21

30 reactions such as dark chemical reduction of Hg (Zheng & Hintelmann 2009, 2010b) and vaporization of liquid Hg to vapor Hg (Estrade et al., 2009). In the next section, Hg isotope fractionation in natural samples as well as experimental studies will be reviewed. 1. MDF of Hg isotopes There are various processes in nature that exhibit only MDF of Hg isotopes in natural systems. Different Hg isotopic compositions observed in upper crust may come from source rocks, hydrothermal systems and ore deposits (Smith et al., 2008). Mass-dependent fractionation has been observed in the source rocks of sedimentary and volcanic nature with δ 202 Hg ranging from +0.2 to -1.0 and -0.2 to -0.6 respectively (Smith et al., 2008). In continental hydrothermal systems, Hg isotopic composition (δ 202 Hg) observed varied from to +2.1 (Smith et al., 2005; Smith et al., 2008; Sherman et al., 2009) without any mass-independent fractionation observed in Hg isotopes. The volatilization of Hg 0 (aq) from magma or country rocks into Hg 0 (g) during boiling reactions or separation of vapor phase resulted in the enrichment of lighter Hg isotopes in vapor phase with low δ 202 Hg values. This vapor phase then oxidized to Hg (II) and reacted with reduced sulfur to form cinnabar precipitates at the surface. Therefore, sedimentary rocks, spring deposits and other surface deposits have a high degree of lighter Hg isotopes while the deep hydrothermal fluid is accumulated with heavier Hg isotopes and high δ 202 Hg values. Sherman et al., (2009) also reported mass-dependent fractionation in marine hydrothermal system (Guaymas basin seafloor rift) with δ 202 Hg ranging between to -0.1 ± (2SD). In addition to MDF, small Δ 199 Hg values of 0.13 ± 0.06 (2SD) were also reported for continental hydrothermal system in yellow Plateau volcanic field that were likely exposed to sunlight (Sherman et al., 22

31 2009). So far, continental and marine hydrothermal systems have been reported to exhibit only MDF of Hg isotopes during volatilization with the exception of yellow Plateau hydrothermal system. Additionally in nature, terrestrial samples such as coal, peat, moss, lichens, soils and sediments and snow and aquatic fish and food web organisms have been observed to display both MDF and MIF. The review of MIF in natural samples is presented in next section. The coal deposits from USA, China and Russia have been investigated for MDF of Hg isotopes. Biswas et al. (2008) observed a range in δ 202 Hg values from -0.1 to Organic soils and sediments such as forest soils, Arctic surficial sediments and organic rich sediments also display a range in MDF values from to -4 (Biswas et al., 2008; Foucher & Hintelmann, 2006; Gantner et al., 2009 and Gehrke et al., 2008). Peat deposits and mosses exhibit δ 202 Hg values up to -1.3 (Biswas et al., 2008; Ghosh et al., 2008) while MDF values in lichens ranges from 1.5 to -2 (Carignan et al., 2009). δ 202 Hg values have also been observed in fish and food web organisms that range from 2 to -3 (Bergquist & Blum, 2007, 2009; Jackson et al., 2008; Laffont et al., 2009; Gantner et al., 2009). Apart from natural samples, the Hg isotope fractionation has also been observed in various experimental studies such as volatilization, biotic and abiotic reduction experiments. The volatilization of Hg 0 from aqueous phase to gas phase in laboratory conditions in presence of tin chloride resulted in MDF with the enrichment of heavier isotopes in Hg 0 aqueous phase (Zheng et al., 2007). The equilibrium and kinetic evaporation of liquid Hg 0 to vapor Hg 0 also displayed MDF (Estrade et al., 2009). Kinetic evaporation displayed large MDF with δ 202 Hg values of ± 0.20 (2SE) while equilibrium evaporation showed small MDF values of 23

32 ~ 0.86 ± 0.22 (2SD). Both kinetic and equilibrium evaporation resulted in isotopically heavier reactants. As such, during the volatilization observed in natural samples and laboratory experiments, heavier isotopes of Hg are preferentially retained in the reactants while lighter isotopes volatilize and accumulate in the vapor Hg 0 phase. The biotic reduction of Hg and MMHg only reported MDF of Hg isotopes (Kritee et al., 2007; 2008; 2009). The microbial reduction of Hg 2+ by various mer-mediated Hg resistant bacteria (i.e. E.coli JM109/pB117, Bacillus cereus 5 and Anoxybacillus Sp. FB9) and non mer-mediated Hg sensitive bacteria resulted in the enrichment of heavier Hg isotopes in the reactant while the trapped Hg 0 was isotopically lighter with low δ 202 Hg values. The massdependent fractionation by 4 unrelated Hg resistant and Hg sensitive bacteria resulted in an overlapping range of fractionation factors with α δ202 from ± (1SD) to ± (1SD) (Kritee et al., 2007, 2008). The microbial reduction of MMHg by mermediated Hg resistant bacteria produced methane and gaseous elemental Hg as products. The mass-dependent fractionation of MMHg by E.coli JM109/pB117 resulted in the reactants being isotopically heavier with α δ202 values of ± (2SD) while the products were depleted in heavier isotopes. The extent of mass-dependent fractionation is lower for MMHg reduction as compared to Hg reduction by various bacteria strains. Mass-dependent fractionation of Hg isotopes has also been observed during the photochemical transformation processes. These processes also display large MIF signatures, discussed in a separate section. The abiotic photochemical reduction of Hg in aqueous systems show MDF in presence of dissolved organic matter (Bergquist & Blum, 2007; Zheng & Hintelmann, 2009) and low molecular weight organic compounds (Zheng & Hintelmann, 24

33 2010a) resulting in the enrichment of heavier isotopes in the reactants while the products were isotopically lighter. Similar Hg (II) photochemical reduction experiments in snow were conducted using flux chamber experiments, where the surface snow samples were isotopically heavier (Sherman et al., 2010). 2. Large MIE MIF Large MIF anomalies for Hg isotopes were first observed during the experimental photochemical reduction of Hg (II) and MMHg under natural sunlight and in fish (Bergquist & Blum, 2007). Large positive MIF signatures ~ 2 were observed during Hg (II) photoreduction in presence of dissolved organic matter, where the odd Hg isotopes were preferentially enriched in the reactants (Bergquist & Blum, 2007; Zheng & Hintelmann 2009; Zheng & Hintelmann, 2010a). It was suggested that the large, positive MIF signatures seen during photo-reduction processes might be a result of magnetic isotope effect. Negative MIF signatures have also been observed in both natural and laboratory experiments involving Hg transformation processes. The photo-reduction of Hg (II) with sulfur containing ligands resulted in significant negative MIF signatures of ~ -1.32, where as the reaction products (Hg 0 ) were enriched in odd Hg isotopes (Zheng & Hintelmann, 2010a). Similar photoreduction experiments in Arctic snow were carried out using flux chamber, which resulted in large negative Δ 199 Hg values up to ± 0.09 (2SD) (Sherman et al., 2010). Following Bergquist & Blum (2007) demonstration of Hg MIF in experimental photoreduction studies, various natural samples such as fish and aquatic food webs (Bergquist & Blum, 2007, 2009; Jackson et al., 2008; Laffont et al., 2009; Gantner et al., 2009; Senn et al., 2010; Gehrke et al., 2010) and terrestrial samples such as coal deposits, peat, moss, lichens, 25

34 surface sediments and snow have been reported to display MIF signatures. The positive MIF signatures in fish range up to +5 in various natural waters. It has been suggested that the positive MIF signatures in aquatic organisms such as fish is due to the MeHg photodegradation in aqueous systems (Bergquist & Blum, 2007, 2009). In contrast, negative MIF signatures have been observed in various terrestrial samples, which may possibly be due atmospherically deposited Hg. The photochemical reduction processes of Hg in aqueous systems produce Hg 0 into the atmosphere, which carries MIF signatures. It has been suggested that these negative MIF signatures may possibly be preserved in the atmosphere and deposited on to the terrestrial surfaces (Bergquist & Blum, 2007). Coal deposits have been investigated for MIF and show depletion in odd Hg isotopes with Δ 199 Hg values up to (Biswas et al., 2008). Similarly, organic surface soils and sediments also report small negative MIF anomalies up to -0.4 in forest soils (Biswas et al., 2008), organic rich sediments from mid-pleistocene Mediterranean Sea (Gehrke et al., 2008) and Arctic lake sediments (Gantner et al., 2009). Peat deposits exhibit very small MIF values up to -0.2 (Biswas et al., 2008). In addition, lichens that derive their nutrients from atmosphere by trapping the atmospheric matter MIF signatures exhibit Δ 199 Hg values ranging between -0.3 to Large MIF signatures were also observed in Arctic snow samples (i.e. surface snow, snowfall and drifted snow). Based on the Hg photochemical reduction experiments with snow, Sherman et al., (2010) interpreted the MIF signatures in natural snow samples to be a result of photo-reduction of Hg where Δ 199 Hg values reached up to ± 0.09 (2SD). This observed negative MIF anomalies in natural samples and laboratory experiments was suggested to be a result of MIE. 26

35 3. Small NVE MIF The abiotic chemical reduction of Hg also exhibits both MDF and MIF. In the dark reduction of Hg in presence of organic matter, small positive Δ 199 Hg values up to 0.4 in the trapped Hg 0 were reported such that odd Hg isotopes were depleted in the reactant (Zheng & Hintelmann, 2009). The MIF observed in dark chemical reduction was suggested to be a result of nuclear volume effect. The isotopic fractionation during vaporization of liquid Hg 0 vapor Hg 0 resulted in both equilibrium and kinetic fractionation. The equilibrium evaporation exhibited small MIF of Hg isotopes with Δ 199 Hg values ranging from 0.10 to 0.14 while kinetic evaporation reported MIF values ranging from 0.08 to 0.05 in vapor phase. Estrade et al., (2009) suggested that nuclear volume effect is responsible for small MIF observed during evaporation and results in the enrichment of odd Hg isotopes in the vapor phase. NVE has been shown to result in small magnitude in MIF of Hg isotopes and the odd Hg isotopes are enriched in the products Δ 199 Hg/Δ 201 Hg slope The MIF of Hg isotopes was first observed during the photochemical reduction of Hg 2+ and MMHg in aqueous solutions in presence of dissolved organic matter. The photo-reduction of Hg 2+ and MMHg in aqueous systems produced two distinct Δ 199 Hg/Δ 201 Hg ratios of 1.03 ± 0.02 (2SE) and 1.36 ± 0.03 (2SE) respectively. Bergquist & Blum (2007) suggested that the relationship between Δ 199 Hg and Δ 201 Hg might possibly be used to distinguish different transformation processes in Hg cycle. If the MIF signatures produced during different Hg transformation processes are unique, then the Δ 199 Hg/Δ 201 Hg ratios could possibly be used identity and trace different photochemical reduction pathways in Hg biogeochemical cycle. 27

36 The photochemical reduction of Hg (II) in presence of dissolved organic matter exhibited a Δ 199 Hg/Δ 201 Hg slope of 1.03 ± 0.02 (2SE) at very high Hg/DOC ratios (Bergquist & Blum, 2007). This ratio is similar to Δ 199 Hg/ Δ 201 Hg ratio observed in Arctic snow samples (Sherman et al., 2010) during Hg (II) photo-reduction. In contrast, Zheng & Hintelmann (2009) observed a range in Δ 199 Hg/Δ 201 Hg ratios from 1.19 ± 0.01 to 1.31 ± 0.07 (1SE) at increasing Hg/DOC ratios. This range in Δ 199 Hg/Δ 201 Hg ratios was attributed to the binding of Hg 2+ to sulfur and oxygen ligands during photo-reduction of Hg. For MMHg photodegradation in presence of DOM, Bergquist & Blum (2007) observed a distinct Δ 199 Hg/Δ 201 Hg slope of 1.36 ± 0.03 (2SE), which is similar to the slope obtained by Rose et al. (in preparation) ~ 1.35 ± 0.16 (2SE) under varying UV radiation and 1.28 ± 0.03 (2SD) in acidic and alkaline solutions (Malinskovy et al., 2010). Therefore, Δ 199 Hg/Δ 201 Hg ratios may possibly be used to identify different Hg and MMHg photo-reduction pathways in nature. As mentioned before, the MIF signatures in the photochemical Hg processes is likely due to MIE. MIF of Hg isotopes can also be induced by nuclear volume effect (NVE). So far, NVE has been observed in equilibrium experiments and abiotic dark reduction experiments. The NVE slope of 1.61 ± 0.06 (2SE) was determined for dark reduction of Hg (II) in presence of DOM and SnCl 2 (Zheng & Hintelmann, 2009, 2010b), 1.54 ± 0.22 (2SE) for equilibrium fractionation between dissolved Hg 2+ and thiol-complexed Hg (Wiederhold et al., 2010) and 1.62 ± 0.06 (2SE) and 2.0 ± 0.6 (2SE) for evaporation experiments between liquid Hg 0 to vapor Hg 0. (Ghosh et al., (in prep.); Estrade et al., 2009). The published NVE slopes from Wiederhold et al. (2010) and Estrade et al. (2009) have a large uncertainty such that MIF slope of MMHg photo-degradation lies within this uncertainty. However, the similar slopes 28

37 of ~ 1.61 ± 0.06 (2SE) by Ghosh et al. (in preparation) and Zheng & Hintelmann, (2010) determined during Hg 0 evaporation experiments and dark Hg 2+ reduction experiments can be considered the likely NVE slope, which is distinct from Δ 199 Hg/Δ 201 Hg slopes obtained for photochemical reduction of Hg and MMHg. As such, Δ 199 Hg/Δ 201 Hg ratios can possibly be used to identify and trace Hg transformation pathways in Hg cycle MIF of MMHg in experimental studies The MIF for MMHg was first observed experimentally during the abiotic photochemical reduction of MMHg under natural sunlight. During the aqueous photo-degradation of MMHg in presence of dissolved organic matter, odd isotopes of Hg i.e. Δ 199 Hg and Δ 201 Hg were preferentially enriched in the reactants and displayed MIF values up to 1.7 ± 0.05 (2SE) and 1.22 ± 0.05 (2SE) respectively (Bergquist & Blum, 2007). In addition, Rose (2010) also investigated the effects of different types and amount of VIS/UV radiation during the photochemical reduction of MMHg to determine the expression and extent of MIF. The photo-reduction of MMHg under full natural sunlight (VIS/UV) resulted in Δ 199 Hg values up to 0.55 ± 0.07 (2SD). However, the blocking of UV-A and UV-B radiation greatly suppressed the MIF despite similar loss of MMHg during all the MMHg photo-degradation experiments. The photo-dissociation of MMHg was also observed in different chemical solutions under Hg vapor lamp (Malinovsky et al., 2010). The acidic solutions with low concentrations of dissolved solids displayed large positive MIF signatures with Δ 199 Hg and Δ 201 Hg values up to 1.5 and 1.3. However, in ascorbic acid solution and alkaline solutions, MIF was suppressed due to the presence of radical scavengers, which inhibit the formation of radical pairs. As such, no significant MIF was observed in alkaline and ascorbic acid solutions (Malinovsky et al., 2010). As mentioned in the above section, the MMHg 29

38 photo-degradation appears to have a distinct Δ 199 Hg/Δ 201 Hg slope of ~ 1.35, which might be used to distinguish between Hg and MMHg photo-reduction processes as well as from NVE induced MIF MIF of MMHg in fish In fish, >90% of the Hg is in the form of MMHg (Bloom et al., 1992). As such, the fractionation of Hg isotopes observed in fish may be dominantly associated with MMHg. MIF of odd Hg isotopes have been observed in aquatic fish and other food web organisms with Δ 199 Hg values up to 5 (Bergquist & Blum, 2007, 2009; Jackson et al., 2008; Laffont et al., 2009; Gantner et al., 2009; Senn et al., 2010; Gehrke et al., 2010). MIF signatures have been observed in fish samples from various freshwater environments such as Arctic char from Arctic lakes with Δ 199 Hg values up to 4.87 (Gantner et al., 2009), Amazonian fish with Δ 199 Hg ~ 0.55 (Laffont et al., 2009), temperate lake fish with Δ 199 Hg ~ 4.64%o (Bergquist & Blum, 2007; 2009; Jackson et al., 2008), from estuarine environments such as San Francisco Bay fish with Δ 199 Hg ~1.55 (Gehrke et al., 2010), Gulf of Mexico coastal fish with Δ 201 Hg ~ 0.4 and marine water fish with Δ 201 Hg ~ 1.5 (Senn et al., 2010). As mentioned previously, the relationship between Δ 199 Hg and Δ 201 Hg has been suggested to be able to identify different transformation pathways and as such, the observed MIF signatures may possibly be used to quantify different Hg transformations in Hg biogeochemical cycle (Bergquist & Blum, 2007). The observation of MIF signatures in fish samples from various freshwater systems have resulted in Δ 199 Hg/Δ 201 Hg slope of ~1.29 ± 0.05 (2SE) (Bergquist & Blum, 2007, 2009; Jackson et al., 2008; Laffont et al., 2009; Gantner et al., 2009). Smaller Δ 199 Hg/Δ 201 Hg slopes have been observed for estuarine and 30

39 marine environments such as 1.26 ± 0.01 (2SD) for San Francisco Bay (Gehrke et al., 2010) and ~ 1.2 for northern Gulf of Mexico (Senn et al., 2010). Based on the similarity between the MIF slopes observed in freshwater fish and the slope for MMHg photo-degradation (1.36 ± 0.03, 2SE) in Bergquist & Blum (2007), the authors suggested that the MIF signatures seen in fish were possibly the result of photo-degradation of MMHg in aqueous systems before its uptake and incorporation in to the food webs. Since photo-degradation of MMHg in natural waters can account up to ~80% of MMHg loss and large MIF signatures are associated with photochemical reduction of MMHg in aqueous systems, it has been suggested that MIF signatures imparted during MMHg photo-demethylation in freshwater systems may be the likely cause of MIF of MMHg observed in fish Quantification of MMHg photo-degradation loss in natural waters Bergquist & Blum (2007) suggested that the MIF signatures observed in fish could be used to estimate the loss of MMHg due to photo-demethylation processes in aqueous systems. The MMHg photo-degradation estimates were calculated by using the MMHg photo-reduction experiment in the presence of 2 mg/l DOM. The isotopic data from this experiment was fitted to the Rayleigh distillation model to obtain a relationship between Δ 201 Hg and fraction of MMHg remaining in the reservoir. This relation was applied to the measured Δ 201 Hg values in fish samples. When this relation was used in the fish samples from freshwater lakes (i.e. Lake Michigan and New England lake), net MMHg photo-degradation losses of ~68 ± 8% (1SD) and ~55 ± 8% (1SD) were calculated respectively (Bergquist & Blum, 2007). Since then, various studies have used the Rayleigh relationships from Bergquist & Blum (2007) including the photo-degradation of MMHg in 20mg/L DOM to quantify MMHg degradation losses. Laffont et al., (2009) calculated an estimated MMHg loss of ~5% in an 31

40 Amazon River basin, which is ~ order of magnitude lower than the MMHg losses estimated by Bergquist & Blum (2007) for their temperate lakes. Gehrke et al., (2010) used the MMHg-20mg/L DOM experiment to quantify the MMHg loss, which was estimated around ~ 10-30% in San Francisco Bay while Senn et al., (2010) calculated ~ 10-20% MMHg degradation in the coastal area of northern Gulf of Mexico and ~ 40-65% in open ocean. The photo-degradation estimates calculated in natural waters (as discussed in MMHg degradation section) are roughly similar to the MMHg photo-degradation losses estimated using Hg isotopes, which suggest promise for this method. While using the Rayleigh relationships obtained from 2mg/L and 20mg/L DOM experiments in Bergquist & Blum (2007) to estimate MMHg photo-reduction losses in freshwater systems, various assumptions have to be made, which are mentioned below. 1. The first assumption is that the laboratory derived fractionation factors are representative of the processes occurring in natural waters. 2. The natural MMHg photo-degradation can be described by one Rayleigh fractionation factor. 3. The MIF seen in fish is due to the photo-degradation of MMHg before its uptake and incorporation into the food web organisms. 4. It is also assumed that regardless of various MDF processes occurring within the aquatic food web, the MIF signatures are preserved in the fish. 5. MIF of MMHg in fish is not induced by any biological processes or any other processes occurring in nature. 32

41 Various environmental factors can affect the MMHg photo-degradation in aqueous systems including VIS/UV radiation, dissolved organic matter, ph and presence of various metal concentrations in freshwater systems. The major environmental factors such as type and amount of solar radiation, different types and concentrations of dissolved organic matter, and other matrix components could affect the magnitude and expression of MIF during photoreduction of MMHg. Therefore, it is important to investigate the effect of varying solar intensities and frequencies and different matrices including organic matter on the MIF and fractionation factors of MMHg during photo-reduction processes. These effects are discussed below Factors affecting MIF of MMHg during photo-degradation in freshwater systems The photo-degradation of MMHg in freshwater systems is considered an important pathway for the removal of MMHg from surface waters. MMHg loss by photodecomposition is an important sink in aqueous systems, which limits its availability to biota for bioaccumulation and bio-magnification in fish and other aquatic food web organisms. Rose et al. (in preparation) studied the effects of different intensities and frequencies of solar radiation on the expression of MIF during MMHg photochemical reduction. It was observed that MIF induced during MMHg photo-reduction was principally driven by UV-B. However, the blocking of UV radiation during photo-degradation greatly suppressed MIF in MMHg. In addition, the effect of different intensities of natural sunlight on MIF was tested during MMHg photo-reduction experiments, carried out in July (high solar intensities) and September. However, it was inconclusive whether MIF of MMHg was larger in July than September due to large uncertainties associated with fractionation factors. The Δ 199 Hg/Δ 201 Hg ratio for all MMHg experiments was determined to be 1.35 ± 0.16 (2SE) 33

42 (Rose et al. in prep.). The effect of different concentrations of DOM (2mg/L & 20mg/L SRFA) on MMHg photochemical reduction was tested by Bergquist & Blum (2007). They observed that the magnitude and extent of MIF was different for varying concentrations of DOM used. Since these experiments were carried out on different days where the VIS/UV intensity and frequency was different, the possibility of two different environmental factors affecting the observed MIF signatures in Bergquist & Blum (2007) experiments is possible and was acknowledged by the authors. The Δ 199 Hg/Δ 201 Hg ratio for different MMHg-DOM experiments was determined to be 1.36 ± 0.03 (2SE), which is very similar to the MIF ratio determined by Rose et al. (in preparation). However, it is still unclear how different the extent and expression of MIF of MMHg would be in the presence of different types and amounts of DOM. Therefore, in this study we investigate how predictable the MMHg MIF signatures will be in presence of varying types and amounts of DOM. In order to achieve this, an artificial source of light (100W Xe lamp) was used in this study that would provide a continuous spectrum of radiation. This was done to isolate the effects of dissolved organic matter on the expression and magnitude of MIF signatures observed in the MMHg-DOM reduction experiments. 34

43 2. Objective of this research study: The objective of this research study was to investigate the isotope fractionation during aqueous photochemical reduction of monomethylmercury (MMHg) by different types and amounts of dissolved organic matter (DOM) in order to assess whether the MIF signatures during MMHg photo-degradation are similar and can be used to identify and quantify photodegradation of MMHg in natural systems. This was achieved by: 1. Studying the effect of MMHg- S red -DOM complexation on the expression and extent of mass-independent fractionation during MMHg photo-reduction. 2. Assessing whether mass-independent fractionation of Hg isotopes in laboratory studies can be used to quantify the MMHg photo-degradation in freshwater systems. 3. Determining if the Δ 199 Hg /Δ 201 Hg slope for MMHg is affected by varying DOM type and amount and whether it is similar to the slope observed for freshwater fish. 35

44 3. METHODS 3.1 Organic Matter: In this study, three different types of organic matter standards were purchased from the International Humic Substance Society (IHSS). The Suwannee River fulvic acid (1R101F), Pony Lake fulvic acid (1R109F) and Nordic Lake natural organic matter (1R108N) were selected based on their elemental sulfur composition in wt%. The elemental analysis of these three organic matter is provided in the Table Sample Preparation: Six MMHg photo-reduction experiments were conducted to study the MMHg-dissolved organic matter (DOM) complexation on photochemical reduction and isotope fractionation. Each type of organic matter was dissolved in 1L Milli-Q water in acid clean glass bottles to obtain a concentration of 2.5mg/L and 20mg/L DOM. MMHg (Alfa Aesar lot A) was added to the 1L organic matter solutions. For each experiment, an initial MMHg concentration of approximately 25µg/L was achieved. The details of the six experiments are provided in Table 3. Once the solutions were prepared, a subsample from every MMHg- DOM solution was taken. To achieve equilibrium between MMHg and dissolved organic matter, the aqueous MMHg-DOM solutions bottles were wrapped in aluminum foil and kept overnight before irradiation by the arc lamp. 3.3 Experimental design: The photochemical reduction of MMHg with DOM was carried out in presence of visible and ultraviolet radiation (UV-A & UV-B) in the Isotope geochemistry lab at University of Toronto. A 100W Xenon arc lamp with an IR filter was used as a source of visible and UV 36

45 radiation. Two quartz-crystallizing dishes (Corning Ware) were used to filter out UV-C radiation from the lamp. The experiments were conducted in 1L quartz Erlenmeyer flasks as shown in Figure 3. The apparatus was designed to pass ambient air through the 0.2µm filter to remove any particulates and a gold trap to remove Hg from ambient air. This design was adapted from Bergquist and Blum (2007). The MMHg-DOM photochemical reduction experiments were carried out for a period of 30 hours. During the photochemical reduction, Hg 0 vapor produced was continuously removed from the flask using a peri-pump at a rate of 250ml/min. Prior to each experiment an initial subsample (40ml) was taken with an acid clean syringe such that initial MMHg concentrations (R0B) and isotopic compositions could be determined. During the experiment, 40ml subsamples were taken from the MMHg-DOM reservoir at various time intervals and preserved with 2.5% BrCl. For every experiment, the irradiation intensity of visible, UV-A and UV-B was measured with a solar light PMA 2200 radiometer. The PAR intensity was measured at 240 W/m Dark controls: The dark control experiments for each organic matter were also carried out for 30 hours in the Isotope geochemistry lab at University of Toronto. The experiments were carried out in aluminum foil wrapped, 1L quartz Erlenmeyer flasks with a DOM concentration of 20mg/L and MMHg concentration of approximately 25µg/L. Prior to each experiment, an initial subsample (40ml) was obtained with an acid clean syringe. During the experiment, subsamples were obtained every 5 hours from the MMHg-DOM reservoir to determine MMHg concentrations and isotope fractionation, if any photo-reduction loss was measured. The samples were then preserved with 2.5% BrCl. 37

46 3.5 Methylmercury analysis: For every MMHg-DOM photochemical reduction experiment, the reservoir subsamples were analyzed by Tekran 2600 cold vapor atomic fluorescence spectrometer (CVAFS). The EPA method 1631, Revision E was used to measure MMHg concentrations. The 2.5% BrCl preserved reservoir subsamples and the trap samples were neutralized by 100% hydroxylamine hydrochloride (HH). These samples were then introduced into the gas-liquid (G/L) separator prior to Hg analysis by CVAFS. Following sample uptake, Hg in the solution was reduced to Hg 0 with 3% tin chloride (SnCl 2 ). The reduced Hg species was captured by 5.0 grade Argon gas, introduced at the bottom of the G/L separator and then pre-concentrated at the dual stage gold traps (Tekran analytical guide: method 1631). The Hg loaded on to the two gold traps was then desorbed and carried to the CVAFS detector. The precision for MMHg concentrations are ± 4.41 %, 1SD, n = 38. The standard reference material NIST1641d was used to monitor the Tekran drift while analyzing MMHg concentrations and trap recoveries. 3.6 DOM matrix purification for isotope analysis: In this study, three different types of organic matter were used with concentrations of 2.5mg/L and 20mg/L DOM. These high concentrations of dissolved organic matter resulted in complex MMHg-DOM matrix. Therefore, to remove sample matrix effects that would affect the Hg isotope compositions (both mass-dependent and mass-independent measurements) (Bergquist & Blum 2009), Hg from the reservoir subsamples was trapped in 1% KMnO 4 and 10% TMG H 2 SO 4, using offline trapping system. In offline trapping, the samples were introduced to the G/L separator along with 10% SnCl 2 for reduction to Hg 0 vapor. This Hg 0 vapor produced was then trapped into 35ml of KMnO 4 - H 2 SO 4 trapping 38

47 solution, where Hg 0 was oxidized. After every sample uptake, the offline system was cleaned with 10% H 2 SO 4 and 10% HNO 3 for minutes. Any Hg 0 sticking to the interior of the trapping Teflon line was also removed and trapped in 35ml trapping solution using a heat gun. If the Hg was not desorbed from the trapping Teflon line, poor sample recoveries were observed. The trapping Teflon line in the acidic permanganate solution was cleaned using 100% hydroxylamine hydrochloride (HH) and 10% H 2 SO 4. NIST 3133 Hg standard was trapped at the beginning of each session to ensure Hg recoveries were over 95%. The concentration and matrix of NIST 3133 was matched to the sample concentration and matrix. All the sample recoveries used for isotope analysis are over 95%. 3.7 Stable isotope analysis: Hg isotopic ratios were measured by multi-collector inductively coupled plasma mass spectrometer (MC-ICP-MS; Thermo scientific) at the Department of Geology, University of Toronto (Figure 4). The matrix purified samples and standards (i.e. trapped samples) were introduced into the mass spectrometer through cold vapor generation system. Prior to the uptake of samples through G/L separator, the trapped samples and standards were neutralized by 100% hydroxylamine hydrochloride (HH) and diluted to a concentration of approximately 5 µg/l. The Hg (II) was reduced to Hg 0 vapor by 10% SnCl 2 on G/L separator and mixed with aerosols of thalium internal standard (NIST 997). These aerosols were produced from Aridius II desolvating nebulizer (CETAC). The Hg 0 and thalium vapor were introduced into the plasma through argon gas, as shown in Figure 4 (Blum & Bergquist, 2007; Bergquist and Blum, 2009) 39

48 The instrumental mass bias was corrected by monitoring the Tl 205 /Tl 203 ratio to determine the instrumental mass bias factor and correct for Hg isotopic compositions. Sample-standard bracketing (SSB) with NIST3133 Hg standard was also used to correct for mass bias. The concentrations of the samples were matched within ±10% of the bracketing standard NIST Also, the matrix of the bracketing standard was matched to the sample matrix of 1% KMnO 4 and 10% H 2 SO 4. In addition, on peak zero correction (i.e. blank correction) was applied to all isotope masses and Pb 204 interference on Hg 204 was corrected by monitoring Pb 206 (Blum & Bergquist, 2007; Kritee et al., 2007 and Bergquist and Blum, 2009). For mass-dependent fractionation (MDF), Hg isotopes were reported in delta notation with per mil ( ) deviations from NIST3133 standard. δ X Hg = [( X Hg/ 198 Hg) sample /( X Hg/ 198 Hg) NIST3133 std ) 1] x 1000 (1) where, x is the mass of Hg isotope between Hg 199 and Hg 204. For mass-independent isotope fractionation (MIF), the Hg isotopes were reported as cap delta notation, in units of per mil ( ). MIF is the deviation in isotopic ratios from the theoretical values predicted by MDF. δ X Hg = δ X Hg - (β x x δ 202 Hg) (2) where, x is the mass of Hg 199, Hg 200, Hg 201, and Hg 204 and β x is the kinetic fractionation factor for that isotope. 40

49 Δ 199 Hg = δ 199 Hg - ( x δ 202 Hg) (3) Δ 200 Hg = δ 200 Hg - ( x δ 202 Hg) (4) Δ 201 Hg = δ 201 Hg - ( x δ 202 Hg) (5) Δ 204 Hg = δ 204 Hg - (1.493 x δ 202 Hg) (6) (Bergquist and Blum, 2007; 2009) 3.8 Analytical uncertainties: Analytical uncertainty during isotope analysis in this study was determined based on measurements of the secondary standard JT Baker. This secondary standard was used relative to NIST 3133 standard, whose isotopic composition is different from the bracketing standard. For JT Baker, δ 202 Hg was measured as ± 0.18 (2SD, n = 24) and average Δ 199 Hg was measured as 0.02 ± 0.06 (2SD, n= 24). The isotope values of were reported using 2SD and analyzed several times during the isotope sessions. 3.9 Calculation of fractionation factors (α): In this study, the kinetic fractionation factors α δ202 and α Δ199 were calculated from the MMHg-DOM isotopic data, which was fitted to the Rayleigh distillation model using Eq (7). The right hand side of the Eq (7) is plotted as a function of ln(f) using Rayleigh distillation equation from Mariotti at al., (1981) to determine α. (10-3 ε p/r )ln(f) = ln(10-3 δ x Hg t +1)/ ln(10-3 δ x Hg 0 +1) (7) where, ε p/r - Enrichment factor ( ) f - fraction of MMHg in the reservoir 41

50 δ x Hg 0 - Initial isotopic composition of Hg where x = 199 Hg to 204 Hg δ x Hg t - Isotopic composition of Hg at time (t) where x = 199 Hg to 204 Hg The slope of best linear fit of relation ln (Δ 199 Hg) or ln (δ 202 Hg) as a function of ln (f) represents the enrichment factor (ε p/r), which was used to calculate kinetic fractionation factors α δ202 and α Δ199. The fractionation factor is expressed as α p/r = 1 + (ε p/r/1000) (8) where, α p/r - kinetic fractionation factor (R products /R reactants ) The uncertainty in α p/r and ε p/r for all MMHg DOM isotopic data is based on 2SE. 42

51 4. RESULTS 4.1. MMHg reduction kinetics Photochemical reduction of MMHg: Monomethylmercury was photo-chemically reduced in 1L Erlenmeyer reactors (as described in methods section) in the presence of different types and amounts of dissolved organic matter. The MMHg loss from the photochemical reduction of MMHg was determined by monitoring the concentrations of MMHg in the MMHg DOM reservoir (obtained through sampling) where the concentration of MMHg decreased with time. When the reservoir subsamples were analyzed through Tekran, THg (total Hg) concentrations were measured. Since the speciation of Hg is unknown as the reaction progressed in the reservoir, it was assumed that MMHg is directly photo-degraded to Hg 0 in the presence of different types and amounts of DOM without any Hg 2+ contamination in the reservoir. Six MMHg-DOM experiments were carried out over a period of 30 hours in presence of an artificial source of light. For 2.5mg/L and 20mg/L of Suwannee River fulvic acid (SRFA) and Pony Lake fulvic acid (PL FAR), the MMHg reduction loss was between 21% to 24% over a period of 30 hours (Figure 6). However, for Nordic Lake natural organic matter (NL NOM), MMHg concentrations showed a loss of 35% for 2.5mg/L and 14% for 20mg/L NL NOM (Figure 6). Dark abiotic reduction of MMHg: The dark abiotic MMHg reduction experiments were only carried out for MMHg -20mg/L DOM for a period of 30 hours. In all the three experiments i.e. 20mg/L SRFA, PL FAR and NL NOM, no MMHg loss was measured (Figure 6). However, we observed an increase in MMHg concentrations in SRFA and PL FAR dark controls at time intervals of 10 and 15 hours and in NL NOM at time intervals of 10, 15 and 30 hours as seen in Figure 6. The 43

52 reason for this observation is unclear. It is possible that when dark controls were run for MMHg concentrations, the drift in Tekran (~ 5-10%) may have resulted in the increased Hg concentrations. Therefore, before publication, these three dark control experiments will be run for concentrations again Mass-dependent fractionation of MMHg-DOM The photochemical reduction of MMHg with DOM exhibited mass-dependent fractionation of even isotopes of Hg. As mentioned in methods section, δ 202 Hg is used to report MDF values. All the photo-reduction experiments of MMHg displayed kinetic isotope fractionation with an enrichment of heavier Hg isotopes in the reservoir while the lighter Hg isotopes were preferentially reduced and lost as Hg 0. The complete isotope data for MMHg- DOM photo-reduction experiments is shown in Table 5. These experiments displayed MDF for the even Hg isotopes as can be seen in Figures (7B, 9B, 11B) where δ 204 Hg is plotted as a function of δ 202 Hg. The data follows MDF as predicted by kinetic mass fractionation laws (the line plotted in the figures). In the MMHg-SRFA experiments, the reservoir Hg became isotopically heavier by ~ 1 over the duration of 30 hours. (Figure 7A). The degree of MDF for both experiments was similar at the same extent of the reaction (f R = 0.8). For MMHg- PL FAR and MMHg- NL NOM experiments, the fraction of MMHg in the reservoir was isotopically enriched in heavier isotopes up to 0.5 over the course of experiments (Figure 9A, 11A). The extent of MDF for MMHg- PL FAR and MMHg- NL NOM experiments was similar at f R = 0.82 and f R = 0.84 respectively. The MMHg- SRFA experiments exhibit larger MDF than MMHg-PL FAR and MMHg-NL NOM experiments at f R = 0.80 as shown in Figure 13A. The dark abiotic MMHg DOM controls did not show MDF since no photochemical reduction of MMHg was observed. 44

53 Rayleigh fractionation: To determine if Rayleigh distillation model can be fitted to the MMHg-DOM isotopic data, δ 202 Hg was plotted as a function of f (fraction of MMHg remaining in the reservoir) as seen in Figure 13A. This plot was fitted to Rayleigh equation (Eq 7) obtained from Mariotti at al. (1981) such that ln (δ 202 Hg) is plotted as a function of ln (f) as shown in Figure 13B. The linear slope of the plot can be used to estimate the enrichment factor (ε), which was used to calculate kinetic fractionation factor α δ202. Table 4 shows the slopes of the plots, its R 2 values for the best linear fit, 95% confidence intervals and the kinetic fractionation factor α δ202 calculated from Eq (8). In Figure 13B, all the MMHg-DOM experiments appear to follow Rayleigh fractionation. The 2.5mg/L and 20mg/L SRFA experiments showed similar fractionation with α δ202 values of ± (2SE) and ± (2SE) respectively. These experiments display a larger extent of fractionation when compared to Pony Lake fulvic acid and Nordic Lake organic matter. As such, at the same extent of reaction, the degree of MDF for both 2.5mg/L and 20mg/L Suwannee River FA experiments is larger than the Pony Lake FAR and Nordic Lake NOM (Figure 13A & 13B and Table 4) Mass-independent fractionation of MMHg-DOM The photochemical reduction of MMHg with different types and amount of DOM also displayed mass-independent fractionation of odd isotopes of Hg i.e. 199 Hg and 201 Hg as can be seen in Figures (8B, 10B, 12B) where δ 199 Hg is plotted as a function of δ 202 Hg. The data shows the deviation of odd Hg isotopes from the mass-dependent fractionation line as predicted by kinetic mass fractionation laws (the line plotted in the figures). The complete isotope data for MMHg-DOM photo-reduction experiments is shown in Table 5. For all the 45

54 MMHg-DOM experiments, significant MIF was observed where Δ 199 Hg and Δ 201 Hg showed positive MIF values for the fraction of MMHg remaining in the reservoir (f) (Figures 8A, 10A & 12A). As the reaction progressed during MMHg-DOM photo-reduction experiments, odd isotopes of Hg were enriched in the MMHg-DOM reservoir and depleted in the evading gas Hg 0. As shown in Figure 8A and 12A, the magnitude of MIF displayed was higher for 20mg/L SRFA and 20mg/L NL NOM than 2.5mg/L SRFA and 2.5mg/L NL NOM at the same extent of reaction at f R = 0.80 and f R = 0.85 respectively. This pattern of MIF for MMHg photo-reduction with SRFA and NL NOM experiments is consistent with the MMHg-SRFA data obtained by Bergquist and Blum (2007). Although the experiments by Bergquist and Blum (2007) were performed on different days with varying solar intensities, the degree of MIF observed for 20mg/L DOM was larger than 2mg/L DOM at the same extent of the reaction (f R = 0.80). In contrast, the degree of MIF for MMHg- PL FAR experiments was similar for both 2.5mg/L and 20mg/L at f R = 0.80 (Figure 10A) Rayleigh fractionation: To determine if Rayleigh distillation model can be fitted to the MMHg-DOM isotopic data, Δ 199 Hg was plotted as a function of f (fraction of MMHg remaining in the reservoir) as seen in Figure 14A. This plot was fitted to Rayleigh equation (Eq 7) obtained from Mariotti at al. (1981) such that ln (Δ 199 Hg) is plotted as a function of ln (f) as shown in Figure 14B. The linear slope of the plot can be used to estimate the enrichment factor (ε), which was used to calculate kinetic fractionation factor α Δ199. Table 4 shows the slopes of the plots, its R 2 values for the best linear fit, 95% confidence intervals and the kinetic fractionation factor α Δ199 calculated from Eq (8). In Figure 14B, MMHg-DOM isotope data followed similar 46

55 MIF slopes with the exception of two experiments that deviated from this data. The 20mg/L SRFA, 20mg/L NL NOM, 2.5mg/L and 20mg/L PL FAR followed similar MIF slopes and are represented as High MIF slope in the Rayleigh plot. These four experiments correspond to low MMHg: S red -DOM ratios of 1:10, 1:21, 1:14 and 1:116 respectively (Table 3). These 4 experiments follow Rayleigh fractionation where α Δ199 calculated are in the range from to and within the error of each other as shown in Table 4. Two experiments that deviate from this data are plotted in Figure 14B that correspond to low DOM concentrations i.e. 2.5mg/L SRFA and 2.5mg/L NL NOM. In addition, the maximum and minimum α values are plotted along each MIF slope in Figure 14B to ensure that we obtain three distinct MIF slopes for different types and amounts of DOM. The mid slope in Figure 14B represents 2.5mg/L SRFA with a high MMHg: S red -DOM ratios of 1:1.5. This experiment did not seem to follow Rayleigh fractionation. In contrast, 2.5mg/L NL NOM experiment follows Rayleigh fractionation and corresponds to low MIF slope with MMHg: S red -DOM ratio of 1:2.4 in Figure 14B. The High MIF slope corresponding to low MMHg: S red -DOM ratios exhibited similar kinetic fractionation factors with an average α Δ199 of ± (2SE). However, as the MMHg: S red -DOM ratios increase, the value of α Δ199 decreases such that high MMHg: S red -DOM experiments showed smaller fractionation with α Δ199 ~ ± (2SE) for 2.5mg/L SRFA and ± (2SE) for 2.5mg/L NL NOM Δ 199 Hg/Δ 201 Hg ratio Bergquist & Blum (2007) suggested that the relationship between Δ 199 Hg and Δ 201 Hg might possibly be used to distinguish different transformation processes in Hg cycle. If the MIF 47

56 signatures produced during different Hg transformation processes are unique, then the Δ 199 Hg/Δ 201 Hg ratios could possibly be used identity and trace different photochemical reduction pathways in Hg biogeochemical cycle. For all MMHg DOM photochemical reduction experiments, the relationship between the MIF of the odd isotopes of Hg, i.e. Δ 199 Hg and Δ 201 Hg, was determined as 1.39 ± 0.06 (2SE) as shown in Figure (15A). This ratio, however, is very different from the MIF slope of photochemical reduction of Hg (II) in aqueous systems (Figure 15B). Bergquist & Blum (2007) obtained Δ 199 Hg/Δ 201 Hg of 1.03 ± 0.02 (2SE) at very high Hg/DOC ratios while Zheng & Hintelmann (2009) obtained a range of 1.19 ± 0.01 (1 SE) to 1.31 ± 0.07 (1SE) with increasing Hg/DOC ratios. 48

57 5. Discussion: Mass-independent fractionation of MMHg-DOM Mechanism of MIF in MMHg photochemical reduction The photochemical reduction of MMHg in the presence of different types and amounts of organic matter resulted in large positive MIF signatures with an enrichment of magnetic Hg isotopes (odd) in the reactants and the depletion of non-magnetic Hg isotopes (even) in the products. The MIF expressed in Hg isotopes during the photo-degradation of MMHg in presence of reduced sulfur ligands is likely a result of magnetic isotope effect (MIE). The MIF observed in the photochemical reduction of Hg (II) and MMHg in presence of dissolved organic matter (Bergquist & Blum, 2007; Zheng & Hintelmann 2009) and photodecomposition of MMHg in acidic and alkaline solutions (Malinovsky et al., 2010) was suggested to be due to MIE, where photo-degradation of MMHg can be expressed by radical pairs as shown in Figure 5. Since both MIE and NVE result in MIF in the odd isotopes, it can be difficult to prove that MIF observed in photo-reduction experiments is due to MIE only. However, there are several arguments that can be made to support this interpretation. 1. The photolytic degradation of MMHg in presence of a ligand results in the formation of radical pair intermediates. Tossell (1998) showed that when MMHg is bound to thiol ligands, its photodecomposition results in the formation of CH 3 and Hg-thiol intermediates (Figure 5). The presence of radical pair intermediates in triplet state is considered a necessary step for MIE to be expressed during photochemistry. But the MIE induced MIF of MMHg can also be suppressed. Malinvosky et al., (2010) was able to suppress MIE in acidic and alkaline 49

58 solutions due to the presence of radical scavengers, which inhibit the formation of radical pairs. 2. The magnitude of MIE induced MIF has been observed to be larger than the MIF produced by NVE. Hg evaporation experiments by Estrade et al. (2009) and Ghosh et al., (in preparation) observed small MIF values up to ~ 0.14 and ~ 0.31 in the vapor phase while kinetic experiments such as abiotic reduction of Hg in presence of DOM (Zheng & Hintelmann, 2009) observed ~ 0.4 in the products. This magnitude of MIF observed by Zheng and Hintelmann, (2009) is one of the largest NVE MIF seen so far in laboratory experiments at a very high extent of reaction. In addition to Hg, the MIF induced by NVE for other elements such as uranium (Fujii et al., 1989a); thalium; zinc, nickel and tellerium (see review: Fujii et al., 2009) is also quite small. 3. The Δ 199 Hg/Δ 201 Hg ratios produced during various Hg transformation processes have been observed to be different for MIE and NVE induced MIF. The Δ 199 Hg/Δ 201 Hg slope of 1.36 ± 0.03 (2SE) for photochemical reduction of MMHg observed by Bergquist and Blum (2007) is within the error of slope determined by Rose et al. (in preparation). However, NVE slope of ~ 1.61 ± 0.06 (2SE) was determined for dark reduction of Hg (II) in presence of DOM and SnCl 2 (Zheng & Hintelmann, 2010b) and liquid-vapor Hg evaporation experiments by Ghosh et al. (in preparation). Zheng and Hintelmann (2010) suggested that the observed dark MIF in kinetic, abiotic Hg (II) reduction reactions is due to NVE. As mentioned in introduction section, MIE is a kinetic mechanism and as such, ruling out NVE from kinetic experiments is difficult. However, any MIF observed in equilibrium reactions should be due to NVE. The published NVE slopes of 1.54 ± 0.22 (2SE) and 2.0 ± 0.6 (2SE) from Wiederhold et al. 50

59 (2010) and Estrade et al. (2009) during equilibrium experiments have a large uncertainty such that MIF slope of MMHg photo-degradation lies within this uncertainty. However, the NVE slope from Zheng & Hintelmann, (2010) during dark kinetic experiments matches quite well with the slope from equilibrium experiments by Ghosh et al. (in preparation) of 1.6 ± 0.05 (2SE). Thus, the experimental Δ 199 Hg/Δ 201 Hg slope of ~1.6 is considered to be the likely NVE slope. Hence, the presence of radical pair intermediates, difference in the magnitude of MIF produced by NVE and MIE and the different Δ 199 Hg/Δ 201 Hg ratios suggest that MIF observed during photo-degradation of MMHg in presence of different types and concentrations of DOM is possibly due magnetic isotope effect. Dark abiotic MMHg degradation The dark MMHg degradation experiments did not result in any reduction in presence of different types and amounts of dissolved organic matter. As such, no isotope fractionation was measured. This is consistent with the dark MMHg control experiments done by Bergquist & Blum (2007) and Zhang & Hsu (2010), who also did not observe any reduction loss. None of the above mentioned studies measured Hg isotopes in MMHg dark reduction experiments since there was no loss and thus no chance for isotope fractionation. In a study by Kritee et al. (2009), dark biotic MMHg degradation by mer-mediated Hg resistant bacteria resulted in MDF of Hg isotopes only. Small increases in MMHg concentrations were observed in SRFA and PL FAR dark controls at time intervals of 10 and 15 hours and in NL NOM at time intervals of 10, 15 and 30 hours as shown in Figure 6. First the small increases in concentration need to be re-measured in order to confirm if they are real. Then dark controls should be run for isotopes to assess 51

60 whether these increases could affect our results for the photochemical loss experiments. Even if these increases are real, it is likely that it would not be associated with significant MIF compared to what we observe in this study. Dark transformations of Hg either result in small MIF with a distinct NVE slope (Zheng & Hintelmann, 2010b; Ghosh et al. (in preparation)) or no MIF at all. Relation between MMHg and S red -DOM ratios The MMHg-DOM photo-reduction MIF data was fitted with a Rayleigh distillation model. Most of the MMHg-DOM experiments display similar fractionation (i.e. α Δ199 ~ ± ) with the exception of two experiments i.e. 2.5mg/L SRFA and 2.5 mg/l NL NOM that display smaller fractionation (Figure 14B). With the limited experiments presented in this thesis, it is hypothesized that the two experiments with different fractionation factors might be related to the concentration of reduced sulfur ligands present in different types and amount of dissolved organic matter. As mentioned in the introduction section, a variety of factors could possibly affect the extent and expression of MIF during MMHg photochemical reduction. So far, the effect of varying solar intensities (Rose et al. in preparation) and different concentrations of DOM (Zheng & Hintelmann, 2009) have been tested on the Hg (II) photo-reduction in aqueous systems and observed that the fractionation factors are dependent on these environmental factors. In addition, Bergquist & Blum (2007) did two MMHg-DOM photo-reduction experiments at different concentrations of DOM and observed different magnitudes of MIF. However, it was unclear to the authors whether the differences in MIF were due to the different amounts of DOM or due to different amounts of solar radiation since the experiments were done at different days in natural sunlight. Therefore, we 52

61 examined the effects of a variety of DOM types and amounts on the expression and extent of MIF during MMHg photo-degradation under artificial source of light such that the effects of DOM are isolated in this study. For all the experiments with low MMHg: S red -DOM ratios i.e. 20mg/L SRFA, 20mg/L NL NOM, 2.5mg/L PL FAR and 20mg/L PL FAR as shown in Figure 14B, the extent of fractionation is very similar and within the error of each other. This may be because these are systems where MMHg concentrations are far lower than the reduced sulfur concentrations. As such, at these low MMHg: S red -DOM ratios, all the MMHg is strongly bound to reduced sulfur ligands, which may explain why all these experiments display similar fractionation factors. In natural aquatic systems, the concentration of reduced sulfur ligands is very high with respect to the concentrations of MMHg. Approximately 70-95% of MMHg is likely to be bound to reduced sulfur ligands in natural waters. As such, low MMHg: S red -DOM experiments may model natural waters in that most of the MMHg is bound to reduced sulfur groups, although the experiments in this study were performed with MMHg concentrations that still far exceed those observed in natural systems. If the low MMHg: S red -DOM experiments are a good representation of nature, then several conclusions can be drawn. 1. In natural systems DOM type and amount may not affect the extent of MIF as long as reduced sulfur ligands are in large excess of MMHg, which is almost always the case. 2. If it is assumed that the isotope fractionation of Hg isotopes in freshwater systems is similar to Hg isotopic data obtained in this study, we suggest that the expression and extent of MIF determined for low MMHg: S red -DOM ratios might be a good representative of natural aquatic systems. Other factors such as solar radiation still need to be considered. 53

62 Although only three types of DOM have been tested in this study, it is hoped that regardless of the type and amount of organic matter present, the similarity in fractionation factors for the low MMHg: S red -DOM ratios supports the use of fractionation factors determined for low MMHg: S red -DOM experiments as a tool to quantify MMHg photo-degradation loss in freshwater systems. The use of specific fractionation factors determined in this study should be cautioned as our experiments were done with artificial light source with ~ 2 times the amount of UV of natural sunlight. Therefore, our study suggest that the use of the fractionation factor from Bergquist & Blum (2007) for 20mg/L DOM experiment performed in natural sunlight might be the best choice available currently. The two experiments that deviate from the MMHg-DOM isotope data are the low DOM concentrations i.e. 2.5mg/L for SRFA and NL NOM (Figure 14B). These experiments correspond to high MMHg: S red -DOM ratios of 1: 1.5 and 1: 2.5. The fractionation factors α Δ199 determined for these two experiments are ± (2SE) and ± (2SE), which decrease with an increase in MMHg: S red -DOM ratios. The decrease in the extent of fractionation for this data set is only seen for high MMHg: S red -DOM ratios, where the MMHg concentrations are relatively higher than the concentrations of reduced sulfur ligands. The reason for the deviation of this data is yet unknown. This occurrence may be because all the MMHg in these experiments may not be completely complexed with reduced sulfur ligands. This could be due to several reasons. One possibility is that a fraction of reduced sulfur ligands might be bound to other metal species and not available to MMHg for binding. Several other metals have high stability constants with thiols including copper, nickel, zinc and lead (Morel & Hering, 1993). We cannot guarantee that our experiments 54

63 were not contaminated with other metals even though we used trace metal clean reagents. The largest unknown possible source of metal contamination is the organic matter standards used in this study. Before publication, several metals will be measured in our reservoir samples. Another possibility is that MMHg is bound to other weaker ligands or anions such as chloride (log k ~ 9.0 to 9.6) and hydroxyl species (log k ~ 4.9 to 5.32) whose stability constants are larger than the stability constants for MMHg-carboxyl binding (log k ~ 1.1 to 3.5) (Alderighi et al., 2003). In order to understand the lower fractionation factors of the two high MMHg: S red -DOM experiments, additional experiments corresponding to different high MMHg: S red -DOM ratios need to be conducted. In addition, a better understanding of the availability of reduced sulfur ligands and its complexation to other metal species at low concentrations of reduced sulfur ligands is required. The 2.5mg/L SRFA experiment did not follow Rayleigh fractionation very well as seen in Figure 14B. With the limited data, the isotope data appears to follow two slopes. The reason for this break within the isotope data is unknown. It may possibly be because of the complexation of MMHg to other weaker ligands or anions in the aqueous solution. Experiments with MMHg and different amounts of SRFA (likely lower DOM concentrations) need to be conducted to confirm how different SRFA concentrations behave with respect to 2.5mg/L SRFA and these experiments need to be redone with more time series points. More experiments with different types of DOM have to be carried out in order to support our hypothesis that the extent and expression of MIF may possibly be dependent on the MMHg: S red -DOM ratios. At low MMHg: S red -DOM ratios, similar fractionation factors are observed 55

64 and as such, it was proposed that when the concentrations of reduced sulfur ligands exceeds the MMHg concentration, the extent and expression of MIF signatures can be used to quantify MMHg photo-reduction loss in freshwater systems. However, as the MMHg: S red - DOM ratios increase, the fractionation factors α Δ199 decrease. For 2.5mg/L SRFA, the highest MMHg: S red -DOM ratio (1:1.2) corresponds to fractionation factor which is greater than the MMHg: S red -DOM ratio (1:2.5) for 2.5mg/L NL NOM (Figure 14B). Therefore, to fully understand the fractionation factors observed in this study for high MMHg: S red ratios, it is important to explore different MMHg: S red -DOM experiments, which would result in MIF fractionation factors that lie above and below the mid MIF slope for 2.5mg/L SRFA. However, the initial isotopic data for MMHg photochemical reduction with different types and concentrations of DOM suggests that the type of organic matter might not be important in nature if there is enough reduced sulfur ligands present in freshwater systems that exceeds the low MMHg concentrations. Natural sunlight vs. lamp data In literature, studies have used natural sunlight and artificial source of light to conduct photoreduction experiments of Hg. The artificial source of light has been used in studies because it provides a continuous spectrum of light intensity that does not change on different days and the ability to isolate the effects of environmental factors such as presence of dissolved organic matter on the expression and magnitude of MIF. However, the isotope fractionation obtained from both artificial sources of light and sunlight needs to be compared to see if the MIF fractionation factors determined in laboratory experiments hold true for natural sunlight as well. As such, the MMHg-DOM MIF data from this study was compared to the MIF data obtained by Bergquist & Blum (2007) (Figure 15). The increase in MMHg: S red -DOM ratios 56

65 observed a decrease in MIF fractionation factors by Bergquist & Blum (2007). This trend in the decrease in fractionation factors was similar to this study. However, the magnitude of MIF fractionation is very different for both sunlight and lamp isotope data. This may be because different amounts of UV can affect the degree of MIF (as shown by Rose et al. 2010), making it difficult to compare lamp data with sunlight data. Relationship between Δ 199 Hg and Δ 201 Hg The Δ 199 Hg and Δ 201 Hg ratios for all six MMHg-DOM photo-reduction experiments show a slope of 1.39 ± 0.06 (2SE) as seen in Figure 16A. Bergquist & Blum (2007) observed a similar Δ 199 Hg/Δ 201 Hg slope of 1.36 ± 0.03 (2SE) for MMHg photo-reduction in presence of DOM, which lies within the error of the slope in this study. However, this ratio is distinct from the MIE induced MIF slope for photochemical reduction of Hg (II) in aqueous systems and NVE induced MIF reactions (mentioned in discussion) (Figure 16B). Bergquist & Blum (2007) obtained Δ 199 Hg/Δ 201 Hg of 1.03 ± 0.02 (2SE) at very high Hg/DOC ratios while Zheng & Hintelmann (2009) observed a range of 1.19 ± 0.01 (1 SE) to 1.31 ± 0.07 (1SE) at increasing Hg/DOC ratios. It has been suggested that Δ 199 Hg/Δ 201 Hg ratios can be used to identify different Hg transformation pathways. As such, using Δ 199 Hg/Δ 201 Hg ratios, different photochemical reduction pathways for Hg (II) and MMHg may be distinguished in Hg biogeochemical cycle (Bergquist & Blum, 2007). From the observed MMHg photodegradation slope of 1.39 ± 0.06 (2SE) in presence of different types and amount of DOM, we suggest that this slope might represent a MIF signature for MMHg photo-reduction pathway. 57

66 It is known that MMHg is a toxic contaminant, which accumulates and biomagnifies in fish and other aquatic food web organisms. Large positive MIF signatures have been observed in aquatic fish and other trophic organisms with Δ 199 Hg values up to 5 (Bergquist & Blum, 2007; Jackson et al., 2008; Laffont et al., 2009; Gantner et al., 2009; Senn et al., 2010; Gehrke et al., 2010). The Δ 199 Hg/Δ 201 Hg ratios have been obtained for fish from different aqueous systems such as 1.30 ± 0.04 (2SE) in temperate lake fish (Bergquist & Blum, 2007; 2009; Jackson et al., 2008), 1.32 ± 0.06 (2SE) in Arctic lake fish (Gantner et al., 2009) and 1.28 ± 0.12 (2SE) in tropical Amazonian fish (Laffont et al., 2009), which are plotted in Figure 16B. This similarity in Δ 199 Hg/Δ 201 Hg slopes of fish and MMHg photo-degradation has been used to suggest that MIF signatures seen in fish are due to the MIF signatures induced during MMHg photo-reduction in aqueous systems. These signatures are then incorporated into the aquatic organisms, where they are preserved through the aquatic food chain despite the MDF processes happening in aquatic food webs (Bergquist & Blum, 2007). The above hypothesis was first suggested when the slope of MMHg photo-degradation was only determined from two experiments conducted by Bergquist & Blum (2007) and when the freshwater fish data was limited. Now, we have a much better determined slope of 1.39 ± 0.06 (2SE) for a variety of DOM types and amounts as obtained in this study and shown in Figure 16A. Moreover, there is much more freshwater fish data available that display a slope of 1.29 ± 0.05 (2SE) as plotted in Figure 16B obtained through regression analysis (Bergquist & Blum, 2007; 2009; Jackson et al., 2008; Gantner et al., 2009; Laffont et al., 2009). These two slopes are slightly different and thus call into question whether the suggested hypothesis by Bergquist & Blum (2007) is indeed true. One big assumption in this hypothesis by Bergquist & Blum (2007) was that the biological processes occurring in the 58

67 food web do not induce MIF. As such, MIF observed in fish is the original MIF signature imparted during photo-reduction of MMHg, which is preserved in fish. However, Jackson et al., (2008) suggested that Hg methylation by microbes in sediments could result in MIF of Hg isotopes, which were observed in forage fish and top predator fish samples in three Canadian lakes. Another study by Das et al. (2009) observed that MIF signatures increase in aquatic organisms as they go up the food chain, suggesting that preservation of MIF signatures in food webs might not occur. They also suggested that the MIF of MMHg was possibly produced within fish during metabolism. However, studies by Kritee et al. (2007; 2008; 2009) and Rodriquez et al. (2008) have demonstrated that Hg transformation by microbes i.e. methylation of Hg (II) or demethylation of MMHg only produces MDF but no MIF. In addition, a recent study by Senn et al. (2010) argued against the observations by Das et al. (2009) who did not see any strong correlations between MIF of Hg isotopes and trophic levels in aquatic food webs, suggesting that MIF of Hg isotopes observed in coastal and oceanic fish from Gulf of Mexico does not occur during the transfer of MMHg from lower to higher trophic levels. They also did not observe any MIF associated with the sediments of Gulf of Mexico. As such, it was suggested that large MIF of MMHg in coastal and oceanic fish is probably due to the photo-degradation of MMHg in coastal and oceanic waters prior to its uptake and incorporation in to the food webs (Senn et al., 2010) where the MIF signatures are preserved. There are several reasons why the laboratory MMHg photo-reduction slope of 1.39 ± 0.06 (2SE) might be different than that observed in fish. One possibility is that the MIF in fish is not due to photo-demethylation prior to its incorporation in the food web as argued by several authors. There may exist another degradation pathway in aqueous systems, which 59

68 may capture the actual MIF signature seen in fish such as a biological mechanism. However, as stated above, there is no actual evidence for biological MIF yet. Another possible reason that the slopes are different is that the slope of fish is not only due to photo-demethylation. A small amount of Hg (II) in fish or perhaps another MIF process could possibly explain the small difference in the Δ 199 Hg/Δ 201 Hg slopes of freshwater fish and MMHg photodegradation. There has been a lot of work on MMHg photo-degradation in natural waters (Sellers et al., 1996; 2001; Krabenhoft et al., 2002; Hammerschmidt & Fitzgerald 2006; Lehnheer & St. Louis 2009; Li et al., 2010) and from these studies it is thought that this is an important mechanism for the removal of MMHg from natural waters accounting for up to ~ 80% of the estimated MMHg loss (Sellers et al., 2001; Hammerschmidt & Fitzgerald 2006; Li et al., 2010). Since this is an important pathway in nature, which also results in large MIF of Hg isotopes, it seems likely that MMHg in aqueous systems would have a MIF signature associated with photo-degradation of MMHg. Perhaps laboratory experiments, which are done at high levels of MMHg, do not truly mimic natural photo-demethylation process. MMHg loss sensitivity to the fractionation factors used Monomethylmercury photo-degradation loss has been calculated in various freshwater systems by using fractionation factors determined from experiments from Bergquist & Blum (2007). In order to assess the sensitivity of estimates of MMHg photochemical loss to different choices of fractionation factors, measured Δ 199 Hg values in fish from various freshwater systems were used with two different fractionation factors to estimate the MMHg photo-degradation loss as shown in Figure 17. The fractionation factors chosen for comparison are from 20mg/L SRFA experiment from this study under artificial radiation and from 20mg/L SRFA experiment from Bergquist & Blum (2007), carried out under natural 60

69 sunlight. Fish from various freshwater systems with a large range in Δ 199 Hg were used to estimate the MMHg loss. The measured MIF values in fish varies from very low Δ 199 Hg values (~ 0.50 ) in Amazon basin (Laffont et al., 2009) to mid Δ 199 Hg values ~ 1.67 in Canadian Arctic lakes (Gantner et al., 2009) to high Δ 199 Hg values ~ 3.4 in Lake Michigan (Bergquist & Blum, 2007). For the Amazon basin where the Δ 199 Hg values for fish were low, the MMHg photodegradation loss estimates were small using both fractionation factors i.e. 5-7% using the natural sunlight fractionation factor and 2-4% using the fractionation factor from artificial radiation experiment. In Canadian Arctic lakes with mid Δ 199 Hg fish values, the MMHg loss estimates ranged from 15-23% using natural sunlight fractionation factor and 9-13% using artificial radiation experiment fractionation factor. In Lake Michigan with high Δ 199 Hg fish values, the estimates of MMHg loss were larger with estimates of 30-40% using natural sunlight fractionation factor and 18-24% using the fractionation factor from artificial radiation experiment. The photo-degradation estimates calculated in natural waters (as discussed in MeHg degradation section) using Bergquist & Blum (2007) Rayleigh fractionation factor is roughly two times the MeHg losses estimated using Rayleigh relation in this study (Figure 17). The estimated MMHg photo-degradation loss is very sensitive to different fractionation factors at mid and high Δ 199 Hg MIF fish values. Overall, the difference in the estimated fraction of MMHg loss increases with increasing Δ 199 Hg values. For example, at low MIF fish values the differences in the fraction of MMHg loss calculated from two different fractionation factors is very small such that the varying fractionation factors does not affect 61

70 the estimated photochemical MMHg loss much. However, in the case of mid and high MIF fish values, the estimated MMHg photo-degradation loss is much more affected by varying fractionation factors. For mid and high Δ 199 Hg fish values, the differences are >5% and >10% respectively. Even with 10-20% differences, MIF estimates of photo-reduction of MMHg could offer an independent method with sufficient precision and accuracy over traditional methods of estimating this loss (i.e. mass balance, direct measurement which is highly variable) if a reasonable range in fractionation factors for natural waters can be identified. The work in this thesis suggests that for many natural waters, different types and amount of DOM may not significantly affect fractionation factors. Thus with proper characterization of UV intensity and penetration, it may be possible to identify representative fractionation factors. Of course, other matrices still need to be explored such as seawater and the difference between Δ 199 Hg/Δ 201 Hg for MMHg photo-degradation and the fish Δ 199 Hg/Δ 201 Hg needs to be resolved. 62

71 6. Conclusion The MMHg demethylation in freshwater systems is a significant pathway for the removal of MMHg from surface waters. It is considered an important sink in aquatic systems that can limit the bioavailability of MMHg to aquatic food webs. Major environmental factors that can affect the MMHg photo-degradation in aqueous systems include different intensity and frequency of VIS/UV radiation and different types and amounts of dissolved organic matter, and other matrices (i.e. metals). Some of these factors have been tested in laboratory experiments to assess their effects on photo-reduction of Hg in aqueous systems (Rose et al. (in preparation); Zheng and Hintelmann, 2009; this study). In this study, we investigated the effects of a variety of DOM types and amounts on the isotope fractionation of Hg isotopes during aqueous photo-reduction of MMHg in order to assess whether the MIF signatures in MMHg in nature are unique and can be used to quantify photo-degradation of MMHg in natural systems. The photochemical reduction of MMHg in the presence of different types and amounts of DOM exhibited both MDF and MIF of Hg isotopes such that the reactants were isotopically heavier and enriched in odd Hg isotopes i.e 199 Hg and 201 Hg. The observation of large, positive MIF signatures in this study is likely due to magnetic isotope effect, which is expressed during radical pair photochemistry. Although, MIF induced by NVE has been observed in kinetic dark reduction experiments (Zheng & Hintelmann, 2009, 2010), this phenomenon was ruled out because of its small magnitude and distinct Δ 199 Hg/Δ 201 Hg ratio from the slope observed in this study. 63

72 The effects of MMHg: S red -DOM ratios on the expression and extent of MIF of MMHg were potentially demonstrated in this study. At low MMHg: S red -DOM ratios, similar fractionation factors α Δ199 were observed (average α Δ199 ~ ± ) while at high MMHg: S red - DOM ratios, smaller fractionation factors of ± and ± were observed. Since low MMHg: S red -DOM experiments correspond to systems with concentrations of reduced sulfur ligands exceeding MMHg concentrations, it is suggested that at these low MMHg:S red -DOM ratios, MMHg is bound to reduced sulfur ligands, which may explain the similar extent of MIF of MMHg for low MMHg:S red -DOM experiments. In addition, the high reduced sulfur concentrations relative to MMHg in these experiments may simulate natural waters. Thus, we suggest that the low MMHg:S red -DOM experiments are possibly a good representative of freshwater systems. As such, the expression and extent of MIF of MMHg in nature may not depend on the type and amount of DOM present in aqueous systems as long as the concentrations of reduced sulfur ligands exceed the MMHg concentrations. Moreover, the similar fractionation factors in low MMHg:S red -DOM experiments suggest that low MMHg:S red -DOM experiments could possibly be used quantitatively to estimate the MMHg photo-degradation loss in freshwater systems. However, for high MMHg: S red -DOM experiments, lower fractionation factors were observed. The reason for these smaller fractionation factors is yet unknown but we hypothesized that in these high MMHg: S red -DOM experiments the MMHg may not all bound to reduced sulfur ligands and instead be partially bound to other weaker organic or inorganic ligands with high stability constants. The photo-reduction of MMHg in presence of different types and amounts of DOM also resulted in a consistent Δ 199 Hg/Δ 201 Hg ratio of 1.39 ± 0.06 (2SE), which is distinct from the 64

73 Hg 2+ photochemical reduction slope and NVE induced MIF slope. The Δ 199 Hg/Δ 201 Hg slope of 1.36 ± 0.03 (2SE) (Bergquist & Blum, 2007) and 1.35± 0.12 (2SE) (Rose et al. in preparation) is similar to the slope obtained in this study. We propose that this slope of ~ 1.39 might be a signature of MMHg degradation pathway and is not affected by different types and amounts of DOM in aqueous solutions. In addition, we observed that the MMHg photo-degradation slope is slightly different from freshwater fish slope of 1.29 ± 0.05 (2SE). The reasons for this are unknown but experimental data needs to be reconciled with natural fish data for Hg MIF to be used quantitatively. 65

74 TABLES: Table 1: Summary of conditional stability constants for MMHg S red -DOM complexation Type of DOM Ligand Log K ph References HA and FA isolates from Fawn lake HA and FA isolates from Fawn lake and Lake Vernon SRFA, IHSS peat HA & Baker brook HA Organic soil from Sweden Organic soils from streams and stream banks in Sweden SRFA, aquatic NOM from wetland & spring lake Summary (soil and aquatic OM) Reduced sulfur ligands Reduced sulfur ligands Reduced sulfur ligands (strong site) 10.7 (Weak site) (strong site) (weak site) 6.5 Hintelmann et al., Hintelmann et al., Amirbahmann et al., 2002 Thiol ligands Qain et al., 2002 Thiol groups Karlsson & Skyllberg et al., 2003 Thiol groups ph independent Khawaja et al., 2010 Thiol ligands Dong et al., 2010 Carboxyl Libich & Rabenstein 1973 Amine Rabenstein et al.,

75 Table 2: The elemental composition of organic matter obtained from IHSS Organic Matter C (wt%) H (wt%) O (wt%) N (wt%) S (wt%) Suwannee River fulvic acid (SRFA) Pony Lake fulvic acid (PLFA) Nordic Lake natural organic matter (NL NOM) nd

76 Table 3: Summary of six MMHg-DOM experiments and MMHg:S red -DOM ratios Experiment DOM DOM conc. MMHg conc. MMHg:S red - DOM ratios mg/l 21.68µg/L 1:1.5 2 SRFA 20 mg/l 24.43µg/L 1: mg/l 23.52µg/L 1:14 4 PLFA 20 mg/l 24.22µg/L 1: mg/l 26.12µg/L 1:2.5 6 NL NOM 20 mg/l 23.97µg/L 1:21 68

77 Table 4: Rayleigh fractionation factors for MMHg-DOM experiments δ 202 Hg Experiment ε = 1000(1-α) R 2 2SE Fractionation factor α 2SE MMHg-2.5mg/L SRFA MMHg-20mg/L SRFA MMHg-2.5mg/L PL FAR MMHg-20mg/L PL FAR MMHg-2.5mg/L NR NOM MMHg-20mg/L NR NOM Δ 199 Hg Experiment ε = 1000(1-α) R 2 2SE Fractionation factor α 2SE MMHg-2.5mg/L SRFA MMHg-20mg/L SRFA MMHg-2.5mg/L PL FAR MMHg-20mg/L PL FAR MMHg-2.5mg/L NR NOM MMHg-20mg/L NR NOM

78 Table 5: Mercury isotope data for MMHg photo-degradation in presence of different types and amounts of DOM under artifical radiation MeHg Conc. (ng/g) n Time f (res) (hours) δ 204 Hg ( ) (avg) δ 202 Hg ( ) (avg) δ 201 Hg ( ) (avg) δ 200 Hg ( ) (avg) δ 199 Hg ( ) (avg) Δ 204 Hg ( ) (avg) Δ 201 Hg ( ) (avg) Δ 200 Hg ( ) (avg) Δ 199 Hg ( ) (avg) Δ199/ Δ201 Avg 2SE Experiment n 2SD 2SD 2SD 2SD 2SD 2SD 2SD 2SD 2SD JTBaker Average R0B MeHg-2.5mg/L SRFA R0B R R R R7* R R MeHg-20mg/L SRFA R0B R3* R4* R R R MeHg-2.5mg/L PL FAR R0B R R R5* R R MeHg-20mg/L PL FAR R0B R R R R R R In this study, 2 standard error (2SE) of the sample replicates are used unless they are smaller than the 2 standard deviation (2SD) of the secondary standard JTBaker. In the case where 2SE of the sample replicates is smaller, uncertainity i.e. 2SD of the JTBaker is applied to all the sample isotope data. 2. * ~ For the samples where n =1, 2SD of the JTBaker is used as the uncertainity in this study.

79 Table 5: Mercury isotope data for MMHg photo-degradation in presence of different types and amounts of DOM under artifical radiation Experiment MeHg Conc. (ng/g) n Time f (res) (hours) n δ 204 Hg ( ) (avg) 2SD δ 202 Hg ( ) (avg) 2SD δ 201 Hg ( ) (avg) 2SD δ 200 Hg ( ) (avg) 2SD δ 199 Hg ( ) (avg) 2SD Δ 204 Hg ( ) (avg) 2SD Δ 201 Hg ( ) (avg) 2SD Δ 200 Hg ( ) (avg) 2SD Δ 199 Hg ( ) (avg) 2SD Δ199/ Δ201 Avg 2SE MeHg-2.5mg/L NL NOM R0B R R R R R R9* MeHg-20mg/L NL NOM R0B R R R R R In this study, 2 standard error (2SE) of the sample replicates are used unless they are smaller than the 2 standard deviation (2SD) of the secondary standard JTBaker. In the case where 2SE of the sample replicates is smaller, uncertainity i.e. 2SD of the JTBaker is applied to all the sample isotope data. 2. * ~ For the samples where n =1, 2SD of the JTBaker is used as the uncertainity in this study.

80 FIGURES: Figure 1: The biogeochemical cycle of mercury in nature (Courtesy of Bergquist, 2008). Natural sources as well as anthropogenic sources emit Hg species into the atmosphere of which elemental Hg 0 has a long residence time of 1 year and is globally distributed. In the atmosphere, Hg undergoes various transformations processes. The atmospheric Hg is deposited in terrestrial and aquatic systems, predominantly as Hg (II) species via wet and dry deposition. In many surface systems, the divalent mercury can be reduced to Hg (0), which then volatilizes to the atmosphere. Under suboxic and anoxic conditions, a small portion of Hg (II) can be converted to more toxic form of MMHg dominantly by bacteria. This MMHg can bioccumulate and biomagnify in aquatic organisms. 72

81 Figure 2: The degradation processes of MMHg in aqueous systems. The MMHg is removed from the natural waters through various processes such as abiotic photo-degradation, abiotic chemical reduction, biological MMHg degradation, and burial in sediments (not shown). Only the photochemical processes in natural waters are known to induce MIF of Hg isotopes. Abiotic chemical reduction and biotic reduction induces MDF only. 73

82 Figure 3: MMHg-DOM photochemical reduction apparatus. The photochemical reduction of MMHg with a variety of DOM types and amounts was carried out in presence of visible and ultraviolet radiation (UV-A & UV-B). Two quartz-crystallizing dishes were used to filter out UV-C radiation from the lamp and a water filter was used to remove longer wavelengths (IR). The experiments were conducted in 1L quartz Erlenmeyer flasks, where the ambient air was passed through the 0.2µm filter to remove any particulates and a gold trap to remove mercury from ambient air. The MMHg-DOM photochemical reduction experiments were carried out for a period of 30 hours. During the photochemical reduction, Hg 0 vapor produced was continuously removed from the flask using a peripump at a rate of 250ml/min. This design is adapted and modified from Bergquist and Blum (2007). 74

83 Figure 4: The gas/liquid phase separator (courtesy of B. Bergquist) and introduction systems to the MC- ICP- MS. 75

84 Figure 5: Reaction scheme for MMHg photo-degradation in presence of different types and amounts of DOM. This figure is adapted and modified from Bergquist and Blum (2009). 76

85 Fraction of MMHg remaining in reservoir (f) MMHg - SRFA mg/L SRFA 20mg/L SRFA 20mg/L Dark control Fraction of MMHg remaining in reservoir (f) MMHg - PL FAR mg/L PL FAR 20mg/L PL FAR 20mg/L PL FAR dark control Fraction of MMHg remaining in reservoir (f) MMHg - NL NOM 0.2 Time (hours) mg/L NL NOM 20mg/L NL NOM 20mg/L NL NOM dark control Figure 6: The aqueous photochemical reduction of MMHg in presence of different types and amounts of dissolved organic matter i.e. Suwannee River fulvic acid, Pony Lake fulvic acid and Nordic Lake natural organic matter results in a 20-35% loss over a period of 30 hours. The Dark controls of MMHg were conducted with only high concentrations of DOM i.e. 20mg/L 77

86 1 MMHg - SRFA A δ 202 Hg ( ) f (fracton of MMHg in reservoir) 2.5mg/L SRFA 20mg/L SRFA 1 B 0 δ 204 Hg ( ) δ 202 Hg ( ) 2.5mg/L SRFA 20mg/L SRFA Figure 7: A- Isotopic composition of δ 202 Hg is plotted as a function of the fraction of MMHg remaining in the MMHg-DOM aqueous reservoir (f) for 2.5mg/L and 20mg/L of SRFA. The errors are represented as ± 0.18 (2SD) for δ 202 Hg. B Isotopic composition of δ 204 Hg is plotted as a function of δ 202 Hg and displays MDF of even isotopes of Hg. The solid line plotted in the lower panel graph is the theoretically predicted MDF based on δ 202 Hg values (Young et al. 2002). 78

87 5 4 MMHg - SRFA A Δ 199 Hg ( ) f (Fracton of MMHg in reservoir) 2.5mg/L SRFA 20mg/L SRFA B δ 199 Hg ( ) Δ 199 Hg Δ 199 Hg δ 202 Hg ( ) 2.5mg/L SRFA 20mg/L SRFA Figure 8: A- Isotopic composition of Δ 199 Hg is plotted as a function of the fraction of MMHg remaining in the MMHg-DOM aqueous reservoir (f) for 2.5mg/L and 20mg/L of SRFA. The errors are represented as ± 0.06 (2SD) for Δ 199 Hg. B Isotopic composition of δ 199 Hg is plotted as a function of δ 202 Hg and displays MIF of odd isotopes of Hg that deviates from the theoretically predicted MDF line. 79

88 0.1 MMHg - Pony Lake FAR A δ 202 Hg ( ) f (fracton of MMHg in reservoir) 2.5mg/L PL FAR 20mg/L PL FAR B δ 204 Hg ( ) δ 202 Hg ( ) 2.5mg/L PL FAR 20mg/L PL FAR Figure 9: A- Isotopic composition of δ 202 Hg is plotted as a function of the fraction of MMHg remaining in the MMHg-DOM aqueous reservoir (f) for 2.5mg/L and 20mg/L of PL FAR. The errors are represented as ± 0.18 (2SD) for δ 202 Hg. B Isotopic composition of δ 204 Hg is plotted as a function of δ 202 Hg and displays MDF of even isotopes of Hg. The solid line plotted in the lower panel graph is the theoretically predicted MDF based on δ 202 Hg values (Young et al. 2002). 80

89 5 4 MMHg - Pony Lake FAR A Δ 199 Hg ( ) f (fracton of MMHg in reservoir) 2.5mg/L PL FAR 20mg/L PL FAR 0.3 δ 199 Hg ( ) Δ 199 Hg B δ 202 Hg ( ) 2.5mg/L PL FAR 20mg/L PL FAR Figure 10: A- Isotopic composition of Δ 199 Hg is plotted as a function of the fraction of MMHg remaining in the MMHg-DOM aqueous reservoir (f) for 2.5mg/L and 20mg/L of PL FAR. The errors are represented as ± 0.06 (2SD) for Δ 199 Hg. B Isotopic composition of δ 199 Hg is plotted as a function of δ 202 Hg and displays MIF of odd isotope of Hg, which deviates from the theoretically predicted MDF line. 81

90 MMHg - Nordic Lake NOM A δ 202 Hg ( ) f (fracton of MMHg in reservoir) 1 B δ 204 Hg ( ) δ 202 Hg ( ) - 3 MeHg- 2.5mg/L NL NOM MeHg- 20mg/L NL NOM Figure 11: A- Isotopic composition of δ 202 Hg is plotted as a function of the fraction of MMHg remaining in the MMHg-DOM aqueous reservoir (f) for 2.5mg/L and 20mg/L of NL NOM. The errors are represented as ± 0.18 (2SD) for δ 202 Hg. B Isotopic composition of δ 204 Hg is plotted as a function of δ 202 Hg and displays MDF of even isotope of Hg. The solid line plotted in the lower panel graph is the theoretically predicted MDF based on δ 202 Hg values (Young et al. 2002). 82

91 4 MMHg - Nordic Lake NOM A 3 Δ 199 Hg ( ) f (Fracton of MMHg in reservoir) B 2 δ 199 Hg ( ) 1 0 Δ 199 Hg δ 202 Hg ( ) MeHg- 2.5mg/L NL NOM MeHg- 20mg/L NL NOM Figure 12: A- Isotopic composition of Δ 199 Hg is plotted as a function of the fraction of MMHg remaining in the MMHg-DOM aqueous reservoir (f) for 2.5mg/L and 20mg/L of NL NOM. The errors are represented as ± 0.06 (2SD) for Δ 199 Hg. B Isotopic composition of δ 199 Hg is plotted as a function of δ 202 Hg and displays MIF of odd isotope of Hg, which deviates from the theoretically predicted MDF line. 83

92 A 0.4 δ202hg ( ) Fraction of MMHg remaining (f) B Ln(δ202Hg) Ln (f) mg/L SRFA 2.5mg/L PL FAR 2.5mg/L NL NOM 20mg/L SRFA 20mg/L PL FAR 20mg/L NL NOM Figure 13: A- Isotopic composition of δ202hg is plotted as a function of the fraction of MMHg remaining in the MMHg-DOM aqueous reservoir (f) for 2.5mg/L and 20mg/L of SRFA, PL FAR and NL NOM. The errors are represented as ± 0.18 (2SD) for δ202hg. B The Rayleigh distillation model was fitted to the MMHg-DOM isotopic data using Rayleigh equation from Mariotti at al. (1981), such that ln (δ202hg) is plotted as a function of ln (f). The slope of the plot can be used to estimate the enrichment factor (ε), which was used to calculate kinetic fractionation factors αδ

93 6 5 A 4 Δ 199 Hg ( ) Fraction of MMHg remaining (f) 6 B 5 α Δ199 = ± Ln(Δ 199 Hg) α Δ199 = ± α Δ199 = ± Ln (f) mg/L SRFA 20mg/L SRFA 2.5mg/L PL FAR 20mg/L PL FAR 2.5mg/L NL NOM 20mg/L NL NOM Figure 14: A- Isotopic composition of Δ 199 Hg is plotted as a function of the fraction of MMHg remaining in the MMHg-DOM aqueous reservoir (f) for 2.5mg/L and 20mg/L of SRFA, PL FAR and NL NOM. The errors are represented as ± 0.06 (2SD) for Δ 199 Hg. B The Rayleigh distillation model was fitted to the MMHg-DOM isotopic data using Rayleigh equation from Mariotti at al. (1981), such that ln (Δ 199 Hg) is plotted as a function of ln (f). The slope of the plot was used to estimate the enrichment factor (ε), which was used to calculate kinetic fractionation factors α Δ199 85

94 6 5 4 α Δ199 = ± Ln(Δ 199 Hg) α Δ199 = ± α Δ199 = ± mg/L SRFA Ln (f) 20mg/L SRFA 2.5mg/L PL FAR 20mg/L PL FAR Figure 15: The fractionation factors α Δ199 in this study are compared to the Bergquist & Blum (2007) fractionation factors obtained for 2mg/L and 20mg/L SRFA. 86