Article pubs.acs.org/est
Mercury Isotope Fractionation during Aqueous Photoreduction of Monomethylmercury in the Presence of Dissolved Organic Matter Priyanka Chandan,† Sanghamitra Ghosh,‡ and Bridget A. Bergquist*,† †
University of Toronto, Department of Earth Sciences, 22 Russell Street, Toronto, Ontario M5S 3B1, Canada Indian Institute of Technology, School of Earth, Ocean and Climate Sciences, Bhubaneswar, Odisha 751 007, India
‡
S Supporting Information *
ABSTRACT: Monomethylmercury (MMHg) is a toxic pollutant that bioaccumulates in aquatic food webs. A major mechanism that limits MMHg uptake by biota is photodemethylation in surface waters. Recently, the extent of mass-independent fractionation (MIF) of Hg isotopes preserved in fish is being used to quantify this MMHg sink. Here, the effects of different types and amounts of DOM on Hg MIF during MMHg photodemethylation were investigated to assess how variable MIF enrichment factors may be with respect to changing DOM binding sites. From experiments conducted with varying amounts of reduced organic sulfur (Sred−DOM), the extent and signature of MIF is likely dependent on whether MMHg is dominantly bound to Sred− DOM. Similar enrichment factors were observed for low MMHg:Sred− DOM experiments, where Sred−DOM was in far excess of MMHg. In contrast, significantly lower and variable enrichment factors were observed for experiments with higher MMHg:Sred−DOM ratios. Additionally the relationship between the two odd Hg isotopes that display MIF (Δ199Hg/Δ201Hg) was consistent for the low MMHg:Sred−DOM experiments, while lower Δ199Hg/ Δ201Hg relationships were observed for the higher MMHg:Sred−DOM experiments. These results suggest that both the extent and signature of MMHg MIF are sensitive to different ligands that bind MMHg in nature.
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INTRODUCTION Mercury (Hg), a highly toxic and globally distributed heavy metal, is released from geochemical reservoirs and anthropogenic sources and cycles between the atmosphere, terrestrial and aquatic systems.1,2 In aqueous ecosystems under suboxic and anoxic conditions, a fraction of Hg is methylated, dominantly by bacterial-mediated processes, to form monomethylmercury (MMHg), which is the form that bioaccumulates in aquatic biota.2−4 In order to assess the risk associated with MMHg, increasing efforts are being made to understand the key processes that control the biogeochemical cycle of MMHg in aqueous systems.2 MMHg demethylation is considered an important pathway for the removal of MMHg in natural waters. MMHg can be demethylated through microbial processes5,6 as well as abiotic reduction processes such as photodegradation7,8 and chemical reduction.9−12 In particular, photodegradation of MMHg is considered one of the major sinks of MMHg in natural waters,8 and several studies have been conducted to understand and quantify MMHg photodegradation in surface waters.8,13−16 Although the estimates of MMHg photodegradation are difficult to make and variable, this process is estimated to account for a large fraction (up to 80%) of the MMHg loss from many natural waters.13,14,16 As a major sink of MMHg, it is necessary to quantify and understand the effects of several key environmental factors on © 2014 American Chemical Society
MMHg photodemethylation in aqueous ecosystems. A number of environmental factors such as light intensity and frequency, dissolved organic matter (DOM) type and amount, photoreactive trace metals, other matrix elements (i.e., chloride, hydroxyl radicals) and pH may affect MMHg photodegradation in natural waters.11,12,14−18 Along with light, the type and amount of DOM is considered a major factor in controlling MMHg photodemethylation.17,19 MMHg in natural waters binds to DOM, which affects its transformation, transport, and bioavailability in the aquatic ecosystems.20,21 MMHg is known to preferentially bind to reduced sulfur ligands, specifically thiol ligands, in soil and aquatic organic matter,22−26 and complexation to different ligands such as sulfide/disulfide and polysulfide complexes is thought to be less significant.26 Thus, it is thought that MMHg mostly forms 1:1 complexes with thiol ligands in many natural waters.25 In many aquatic systems, the concentration of MMHg is very low with respect to the concentration of reduced S ligands in DOM resulting in MMHg being dominantly bound to reduced S ligands.21 However, when concentrations of MMHg are higher or other weaker ligands are present in high concentrations, a significant Received: Revised: Accepted: Published: 259
July 16, 2014 November 12, 2014 November 17, 2014 November 17, 2014 dx.doi.org/10.1021/es5034553 | Environ. Sci. Technol. 2015, 49, 259−267
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natural conditions. In Jimenez-Moreno et al.46 significant photodemethylation was likely occurring in experiments carried out in high chloride solutions under visible light. However, these experiments did not display MIF, which may have been due to the complexation by chloride and/or the lack of higher energy UV radiation. Although little is known about how natural variables will affect Hg isotope fractionation during MMHg photodegradation, there are studies showing that DOM type and amount affects the extent and expression of MIF for Hg2+ photoreduction.27,30,47 Thus, it is likely that DOM type and amount may also affect MIF during MMHg photodemethylation. For example, Bergquist and Blum27 observed different magnitudes of MIF for photodemethylation for two concentrations of DOM under natural sunlight. However, these experiments were carried out on different days and it was unclear whether the variability in MIF was due to different DOM concentrations or varying solar radiation. Therefore, before Hg MIF can be used quantitatively to estimate MMHg photodegradation, it is necessary to investigate the effects of environmental variables on the extent and expression of MIF during MMHg photodegradation. Here, we investigated Hg isotope fractionation during aqueous photoreduction of MMHg in the presence of different types and amounts of DOM. In particular, we assessed the effects of varying the organic reduced S (Sred− DOM) on the variability in enrichment factors and in the MIF signature (Δ199Hg/Δ201Hg ratio).
portion of MMHg could also be bound to weaker ligands in the DOM (i.e., O/N functional groups) or in natural waters (i.e., Cl−, OH−).23,25,26 Stable Hg isotope fractionation is an important tool to trace, distinguish and possibly quantify various sources, sinks and transformation pathways in the Hg biogeochemical cycle.27−29 Recent laboratory photoreduction experiments of Hg in the presence of DOM displayed both mass-dependent (MDF) and large mass-independent fractionation (MIF) of odd Hg isotopes (199Hg and 201Hg) suggesting that natural fractionation of Hg isotopes might be useful in both understanding and quantifying Hg photoreduction pathways in natural waters.27 Large MIF (>0.5‰) was observed during the photochemical reduction of both Hg2+ and MMHg in the presence of DOM,27,30 where MIF is likely driven by the magnetic isotope effect (MIE) that can be expressed during radical pair photochemistry.31 It has also been suggested that the relationship between the MIF for the two odd isotopes (Δ199Hg/Δ201Hg) may potentially be used to distinguish different transformation processes in the Hg cycle based on the differences observed in the Δ199Hg/Δ201Hg for MMHg photodegradation (1.36 ± 0.02, 2SE),27 Hg2+ photoreduction (1.00 ± 0.02, 2SE),27 and dark abiotic processes displaying nuclear volume driven MIF (1.59 ± 0.05, 2SE).32−34 Large positive MIF signatures are observed in fish and other aquatic organisms with Δ199Hg values up to 5‰ from various ecosystems.27,35−42 In fish, >90% of the Hg is generally in the form of MMHg43 and variations in Hg isotopes observed in fish are likely dominantly associated with MMHg. In their initial study, Bergquist and Blum27 observed similar Δ199Hg/Δ201Hg ratios for MMHg photodegradation experiments and freshwater fish. It was argued by the authors that (1) MIF signatures in fish were possibly a result of photodemethylation of MMHg in freshwaters prior to uptake and incorporation into fish and other organisms along the aquatic food chain and (2) the extent of MIF observed in fish could possibly be used to estimate the loss of MMHg due to photodemethylation in aqueous systems. Based on the above arguments, several studies have used Hg MIF signatures preserved in fish to estimate MMHg photodegradation in natural waters by using MIF preserved in fish and a Rayleigh model using the experimental enrichment factors estimated in Bergquist and Blum.27 For example, estimates of net MMHg photodemethylation loss using Hg MIF in fish have been derived for several lake ecosystems (∼5− 70%),27,28,35−37,42 for coastal ecosystems (10−30%)39,40,44 and for the open ocean (40−80%).40,41 Currently the enrichment factors derived from the MMHg photoreduction experiments of Bergquist and Blum27 are being used, despite these original experiments being performed with only one type of DOM, at unrealistic MMHg/DOM ratios, and in MQ water devoid of any other matrix elements other than the added DOM. As mentioned above, several environmental factors can affect the MMHg photodegradation in natural waters. However, little is known of how these environmental factors will affect Hg isotope fractionation during photodemethylation. Two Hg isotope studies beyond Bergquist and Blum27 shed some light on isotope fractionation of Hg during photodemethylation, but neither systematically investigated how natural variables affect MIF enrichment factors.45,46 Malinovsky et al.45 investigated the photodissociation of MMHg in different matrix solutions under UV−C and was able to suppress MIF by addition of radical scavengers. Although this study proved the importance of radical pairs in expressing MIF, it is not easily comparable to
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MATERIALS AND METHODS Materials. In this study, three different types of organic matter standards were utilized from the International Humic Substance Society (IHSS): (1) Suwannee River fulvic acid (SRFA, 1R101F), (2) Pony Lake fulvic acid (PL FAR, 1R109F), and (3) Nordic Lake natural organic matter (NL NOM, 1R108N). DOM materials were selected based on different total elemental S contents of 0.44% for SRFA, 0.65% for NL NOM and 3.03% for PL FAR (for more details on the organic matter standards see SI).48 All the acids used in this study were trace metal or Optima grade quality and all the water used was Millipore deionized water (MQ; 18MΩ). Glassware was cleaned with both 10% HCl and 2.5% BrCl followed by thorough rinsing with MQ water. Experimental Design. The organic matter standards were dissolved in MQ water to obtain concentrations between 1 mg/ L and 20 mg/L, and MMHgCl (Alfa Aesar, lot 82−082844A) was added to the organic matter solutions. For each experiment, an initial MMHg concentration of approximately 25 μg/L was achieved (SI Table S1). To achieve equilibrium between MMHg and DOM, the aqueous MMHg−DOM solutions bottles were wrapped in aluminum foil and kept overnight before the photoreduction experiments were performed.27 Photochemical experiments were carried out in quartz Erlenmeyer flasks, similar to those in Bergquist and Blum,27 except an artificial light source was used instead of natural sunlight for 30 h. Briefly, photoreduction experiments were carried out in the presence of visible and ultraviolet radiation (UV−A & UV−B) with a 100W Oriel research Xenon arc lamp with an IR filter (Newport; model# 66905) and two Pyrex filters (Corning Ware) to filter out UV−C. Prior to each experiment, an initial subsample was taken so that the initial Hg concentration and isotopic composition could be determined. During the experiment, subsamples were taken from the aqueous MMHg reservoir at various time intervals and 260
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the run. All the isotope values and errors are reported following the recommendations in Blum and Bergquist.49 MDF compositions of Hg are reported in delta notation (δ) as per mil (‰) deviations from NIST 3133 standard (eq 1).
preserved with 2.5% BrCl for Hg concentrations and isotope analyses. The irradiation intensity of PAR (∼300 W/m2), UV− A (∼120 W/m2) and UV−B (∼3.8 W/m2) was measured with a solar light PMA 2200 radiometer for every experiment. For each organic matter standard, a dark abiotic MMHg reduction experiment with 20 mg/L DOM was also conducted for 30 h with the same experimental setup as the photodemethylation experiments except the quartz Erlenmeyer flask was covered with aluminum foil. In all the dark experiments, no loss of Hg was measured (SI Figure S2). Since the Hg speciation was not determined during the experiments as the reaction progressed in the reservoir, it was assumed that MMHg photodegraded to Hg0 without significant Hg2+ contamination in the reservoir. This is also supported by the mostly high Δ199Hg/Δ201Hg slopes for the MMHg experiments (1.34 ± 0.04, 2SE, n=8), which likely would have been significantly lowered if significant amounts of Hg2+ were in the reservoir undergoing photoreduction as Hg2+ photoreduction has a slope of 1.00 ± 0.02 (2SE).27 Mercury Concentration and Isotope Analyses. Total Hg concentrations in solutions were analyzed by a Tekran 2600 cold vapor atomic fluorescence spectrometer (CVAFS) following EPA method 1631, Revision E. The standard reference material NIST1641d was used to monitor accuracy and precision, and the error for Hg concentrations was ±4.8% (1SD, n=65). To avoid matrix issues for isotope analysis, all the samples were separated from their matrices using a purge and trap system. Samples were neutralized with 30% hydroxylamine-hydrochloride (HH) to neutralize the KMnO4 matrix and then introduced into a gas−liquid (G/L) separator along with 10% SnCl2 for reduction to Hg0 vapor. The Hg0 vapor was trapped into an oxidizing 1.5% KMnO4 −10% H2SO4 solution. Any Hg0 sticking to the interior of the trapping Teflon line was also removed and trapped using a heat gun. If Hg was not desorbed from the trapping Teflon line, poor sample recoveries were observed. Matrix matched NIST 3133 Hg standard was trapped at the beginning of each session to ensure Hg recoveries were over 95%, and blanks were run periodically to ensure negligible Hg carry over. The recovery of every sample and standard was measured, and all sample recoveries used for isotope analysis were over 95%. Stable Hg Isotope Analysis. Hg isotope ratios were measured by cold vapor multicollector inductively coupled plasma mass spectrometer (CV-MC-ICP-MS) on a Thermo Fischer Neptune Plus at the University of Toronto. Prior to analysis, the preserved trap samples and standards were neutralized by HH and diluted to a concentration of 5 μg/L of Hg. The Hg was reduced to Hg0 vapor by 10% SnCl2, separated on a frosted tip G/L separator and carried to the plasma with Argon gas containing a thallium (Tl) internal standard (NIST SRM 997). Thallium was introduced to the G/ L separator as an aerosol using an Aridius II desolvating nebulizer (CETAC). Instrumental mass bias was corrected using both the Tl internal standard and strict sample-standard bracketing (NIST SRM 3133). The matrices and concentrations of the samples and the bracketing standard were matched such that signal intensities of the samples and bracketing standards were within ±10%. Blank corrections were applied to all Hg isotopes and the 204Pb interference on 204 Hg, which was always negligible, was corrected by monitoring 206Pb.27,49 Blank corrections on Tl were not possible because Tl was continuously streamed throughout
δ X Hg = [(X Hg/198 Hg)sample /( XHg/198Hg)NIST3133Std ) − 1] (1)
x1000 199
204
where, X is the mass of Hg isotope between Hg and Hg. The 202Hg/198Hg ratio (i.e., δ202Hg) is used to report MDF.49 MIF isotopic compositions of Hg are reported with capital delta notation (Δ), which is the deviation in the isotope ratios from the theoretically predicted isotope ratios by MDF (eqs 2 and 3). MIF is reported in units of per mil (‰). Δ199Hg = δ199Hg − (β199 × δ 202 Hg)
(2)
Δ201Hg = δ 201Hg − (β201 × δ 202 Hg)
(3)
Only odd isotope MIF was observed in this study and was calculated using β199 = 0.2520 and β201 = 0.7520 estimated from kinetic MDF laws.49 We use Δ199Hg to report MIF in this study. External reproducibility was determined by measurement of secondary standards, UT-JT Baker and ETH Fluka. For the UT-JT Baker standard, we obtained δ202Hg of −0.62 ± 0.13‰, Δ199Hg of 0.02 ± 0.04‰ and Δ201Hg of 0.00 ± 0.07‰ (2SD, n = 44). For the ETH Fluka standard (from Dr. Jan G. Wiederhold, Institute of Geochemistry and Petrology, Zurich), we measured δ202Hg = −1.43 ± 0.18‰, Δ199Hg = 0.09 ± 0.05‰ and Δ201Hg = 0.09 ± 0.03‰ (2SD, n = 12), which agree well with reported values in Jiskra et al.50 Samples were measured in duplicate and errors are reported as 2SE unless the 2SE is smaller than the external reproducibility of the method as estimated by the JT Baker standard. In that case, the 2SD errors of JT Baker standard are assigned as the error of the sample.49 All isotope data is reported in SI Table S2. Mathematical Calculations. Details of both enrichment factor and pseudo-first order rate constant calculations can be found in the SI. Briefly, enrichment factors represent the isotopic difference between the reactant and the instantaneous product in a chemical reaction. Enrichment factors (εδ202, εΔ199) were estimated using a Rayleigh distillation model where, εp/r is the enrichment factor between product p and reactant r and is reported in per mil (‰).51 The slope of the best linear fit of relation between ln(Δ199Hg) or ln(δ202Hg) and ln(f) represents the εp/r.51 The pseudo-first-order rate constants for all the MMHg−DOM photoreduction experiments were calculated from the integrated form of first-order rate law.52 To make comparisons with published first-order rate constants in literature, the PAR normalized degradation rate constants were estimated using PAR intensity of 300 W/m2. In addition, Δ199Hg/Δ201Hg slopes were determined from the best linear fit of the relationship of Δ199Hg and Δ201Hg, and the errors associated with the slope are 2SE as reported in SI Table 4. Estimation of MMHg:Sred−DOM Ratios. The total elemental organic sulfur (from IHSS),48 estimated fraction of reduced sulfur (Sred)20,21,53 for SRFA, PL FAR and NL NOM and their calculated molar concentrations are given in SI Table S1. MMHg is known to form strong 1:1 complexes with thiols in DOM. However, the estimates of Sred in the organic matter standard materials reported in the literature do not distinguish between thiols and other reduced sulfur ligands, which include ligands that form weaker bonds with MMHg.20,53,54 In our 261
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estimates, the molar concentration of Sred in each DOM was estimated based on the reported Sred content even though not all of this is thiols. We recognize the limitation of these estimates and acknowledge that our estimates are overestimations.55 This is discussed in more detail in the Discussion section. However, the estimates are useful as an indication of the relative amount of reduced organic S ligands and results can be compared especially for experiments using varying amounts of the same organic matter. The reduced sulfur fraction of the total organic S was estimated to be ∼35% and ∼59% for SRFA and PL FAR, respectively.20,53 Due to the absence of measurements of Sred for NL NOM, we assumed the fraction of Sred to be 50%.
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RESULTS AND DISCUSSION Hg Isotope Fractionation During Photodegradation Experiments. MMHg−DOM photoreduction experiments resulted in 10−35% loss of Hg in the reservoir (SI Figure S1). Estimated pseudo-first order rate constants calculated were similar for all the experiments (1.3 × 10−3 to 3.6 × 10−3 E−1m2) with no discernible trends and are comparable to recently published rates of 1 × 10−3 to 4.3 × 10−3 E−1m2 for photodemethylation in the presence of different types of DOM.12,19 All the photodegradation experiments exhibited kinetic MDF of even Hg isotopes with an enrichment of heavier Hg isotopes in the reservoir while the lighter Hg isotopes were preferentially reduced and lost as Hg0. Over the duration of the experiments, the reservoir Hg became ∼1‰ heavier for the SRFA experiments and ∼0.5‰ heavier for the PL FAR and NL NOM experiments. The MDF data in the experiments followed Rayleigh fractionation allowing enrichment factors to be estimated. Similar to the rate constants, MDF enrichment factors (εδ202) also did not vary much in our experiments (all but one experiment fall between −1.13 and −2.24‰) and show no discernible trends. This will be discussed below in the section on demethylation mechanisms. All rate constants and enrichment factors (εδ202) are reported in SI Table S3 and S5. In contrast, the MIF results of our experiments do show trends with our estimates of MMHg:Sred−DOM ratios (Figure 2). All the MMHg−DOM photodegradation experiments displayed MIF of the odd isotopes of Hg (199Hg and 201Hg) with preferential enrichment (+MIF) of odd isotopes in the reservoir. The magnitude of MIF was significantly higher and similar for the 20 mg/L SRFA, 2.5 mg/L PL FAR, 20 mg/L PL FAR and 20 mg/L NL NOM experiments at the same extent of reaction (f R = 0.85). These experiments correspond to low MMHg:Sred−DOM ratios of 1:10, 1:14, 1:116 and 1:21 respectively. In contrast, the experiments corresponding to higher MMHg:Sred−DOM ratios (i.e., 1 mg/L, 2.5 mg/L, 5 mg/L, 10 mg/L SRFA and 2.5 mg/L NL NOM) displayed lower magnitudes of MIF at the same extent of reaction ( f R = 0.85). All the MIF data for the MMHg−DOM experiments followed Rayleigh fractionation (Figure 1) and the enrichment factors εΔ199 values are plotted in Figure 2a and are reported in SI Table S4. Relationship Between MIF and MMHg:Sred−DOM Ratios. The MMHg−DOM experiments corresponding to low MMHg:Sred−DOM ratios had similar εΔ199 values that ranged from −13.4‰ to −16.2‰ with an average εΔ199 of −14.9 ± 1.2‰ (2SE, n = 4) (Figures 1 and 2a). In the low MMHg:Sred−DOM experiments, MMHg concentrations are far lower than the reduced organic S ligand concentrations. At these low MMHg:Sred−DOM ratios, the MMHg is likely
Figure 1. Ln(Δ199Hg) is plotted as a function of the fraction of MMHg remaining in the reservoir (f) for all experiments along with linear regressions to different experiments. MIF enrichment factors (εΔ199, ‰) were determined from the slope of the best linear fit between ln(Δ199Hg) and ln(f) using a Rayleigh distillation model.51 See SI for details. Enrichment factors represent the isotopic difference between the reactant and instantaneous product in chemical reactions. The low MMHg:Sred−DOM experiments had an average εΔ199 of −14.9 ± 1.2‰ (2SE, n = 4).
dominantly bound to the reduced organic S ligands, which may explain why the low MMHg:Sred−DOM experiments display similar enrichment factors. The experiments corresponding to high MMHg:Sred−DOM ratios display significantly lower and more variable enrichment factors (Figures 1 and 2a). The reason for the high MMHg:Sred−DOM experiments to display smaller enrichment factors is not completely understood, but is likely related to MMHg not being mostly bound to reduced organic S ligands. In these experiments, the MMHg concentrations are relatively higher and are closer to or exceed the estimated concentrations of reduced organic S ligands. As such, a fraction of MMHg is likely bound to other ligands. Although in some of the experiments it appears that reduced organic S ligands are in excess based on our estimates of Sred−DOM ligands, this is probably not the case as our calculations of available Sred− DOM ligands are overestimates.55 As discussed in the methods, not all the reduced sulfur ligands quantified by the spectroscopic studies20,23,26,53 may be available for MMHg binding as the estimates include several reduced S ligands that have much lower binding affinities for MMHg than thiols.26,54 It is also possible that MMHg is bound to other weaker ligands that may be present in high enough concentrations to compete with the reduced organic S ligands. For example, in our highest MMHg:Sred−DOM ratio experiments, there may have been enough chloride present in the reservoir from the MMHgCl standard to potentially have 10−30% of MMHg bound to chloride. The 10−30% estimates is an upper limit as it was estimated using a low published stability constant for MMHgRSH (log K = 9.11)24 and a high published stability constant for MMHg−Cl (log K = 5.5).56 However, if the average log K values for MMHg−RSH and MMHg−Cl are used,24,25,56,57 estimates of MMHgCl complexes present during the MMHg photodegradation experiments are small and mostly negligible. Given that we really do not know how much of reduced organic S ligand is available for MMHg complexation and that potentially other ligands are present that could bind MMHg, 262
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DOM ratios. For PL FAR, which had the highest Sred, no drop in enrichment factors was observed because the Sred was always in large excess of the MMHg. Although only three types of DOM were tested in this study, the similarity in enrichment factors for the low MMHg:Sred−DOM ratios supports the use of enrichment factors determined for low MMHg:Sred−DOM experiments for natural systems where reduced organic S concentrations far exceed the MMHg concentrations. MIF Signatures (Δ199Hg and Δ201Hg) and Mechanism of MIF. In order to discuss how the MIF data in this study is related to the pathways of photodemethylation and to utilize the relationship between Δ199Hg and Δ201Hg, the two causes of even−odd Hg MIF will be briefly reviewed: the nuclear volume effect (NVE)58,59 and the magnetic isotope effect (MIE).31 Although both the MIE and the NVE result in odd isotope MIF, consistent with the MIF observed in our experiments, MIE is the likely mechanism in our experiments because both the slopes (Δ199Hg/Δ201Hg) and magnitude of MIF (up to 5.5‰) observed are not consistent with the NVE. The NVE is a result of nuclear volumes and nuclear charge radii for large isotopes not being linearly dependent on the number of neutrons and mass.58,59 Differences in nuclear charge radii affect the bond strengths and therefore can result in isotope fractionation. Because the NVE is dependent on nuclear charge radii and these are constant, the NVE results in a Δ199Hg/ Δ201Hg of 1.59 ± 0.05 (2SE),33 which is confirmed both experimentally32,33 and theoretically.34 The NVE Δ199Hg/ Δ201Hg is higher and distinguishable from the Δ199Hg/Δ201Hg slopes observed in photochemical experiments, which have variable and lower slopes.27,30 Another key difference between NVE and MIE is the extent of MIF observed. Generally, MIE results in much larger MIF than the NVE induced MIF (mostly
