Mercury Methylation Rates for Geochemically Relevant HgII Species

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Mercury Methylation Rates for Geochemically Relevant HgII Species in Sediments Sofi Jonsson,†,‡ Ulf Skyllberg,§ Mats B. Nilsson,§ Per-Olof Westlund,† Andrey Shchukarev,† Erik Lundberg,‡ and Erik Björn*,† †

Department of Chemistry, Umeå University, SE-901 87 Umeå, Sweden Umeå Marine Sciences Centre, Umeå University, SE-910 20 Hörnefors, Sweden § Department of Forest Ecology and Management, Swedish University of Agricultural Sciences, SE-901 83 Umeå, Sweden ‡

S Supporting Information *

ABSTRACT: Monomethylmercury (MeHg) in fish from freshwater, estuarine, and marine environments is a major global environmental issue. Mercury levels in biota are mainly controlled by the methylation of inorganic mercuric mercury (HgII) to MeHg in water, sediments, and soils. There is, however, a knowledge gap concerning the mechanisms and rates of methylation of specific geochemical HgII species. Such information is crucial for a better understanding of variations in MeHg concentrations among ecosystems and, in particular, for predicting the outcome of currently proposed measures to mitigate mercury emissions and reduce MeHg concentrations in fish. To fill this knowledge gap we propose an experimental approach using HgII isotope tracers, with defined and geochemically important adsorbed and solid HgII forms in sediments, to study MeHg formation. We report HgII methylation rate constants, km, in estuarine sediments which span over 2 orders of magnitude depending on chemical form of added tracer: metacinnabar (β-201HgS(s)) < cinnabar (α-199HgS(s)) < HgII reacted with mackinawite (≡FeS-202HgII) < HgII bonded to natural organic matter (NOM-196HgII) < a typical aqueous tracer (198Hg(NO3)2(aq)). We conclude that a combination of thermodynamic and kinetic effects of HgII solid-phase dissolution and surface desorption control the HgII methylation rate in sediments and cause the large observed differences in km-values. The selection of relevant solid-phase and surface-adsorbed HgII tracers will therefore be crucial to achieving biogeochemically accurate estimates of ambient HgII methylation rates.

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INTRODUCTION The spread of mercury (Hg) in aquatic and terrestrial ecosystems constitutes significant threats to wildlife1,2 and human health3 as well as to socioeconomy4 worldwide. The most severe effects are caused by the formation and bioaccumulation of monomethylmercury (MeHg, referring to all possible species of methylmercury) formed from inorganic mercuric mercury (HgII, referring to all possible species of inorganic mercuric mercury), primarily under reducing conditions in sediments and soils via the activity of sulfateand iron-reducing bacteria (SRB and IRB, respectively).5,6 The formation of MeHg is a complex, far from fully understood, biogeochemical process driven by factors that control the activity of methylating bacteria, such as the availability of metabolic electron donors and acceptors, and the availability of aqueous phase HgII complexes.7,8 Recent studies have demonstrated a positive correlation between the total concentration of HgII in sediment pore waters and HgII methylation rates,9 whereas in other studies, specific molecules such as neutral HgII sulfides10 or lowmolecular-mass HgII thiols11,12 have been emphasized as © 2012 American Chemical Society

particularly bioavailable. Other recent studies have proposed nanoparticulate HgS(s) as important sources for MeHg formation in bacteria culture experiments.13,14 Although the composition and concentration of aqueous HgII complexes may be important for the uptake of HgII by methylating bacteria, kinetic and equilibrium processes of HgII in adsorbed and solid phases generally control the total concentration of dissolved HgII in sediment and soil pore waters.9,15 The major adsorbed and solid forms of ambient HgII in anoxic sediments include complexes with thiols in natural organic matter (NOM),16 complexes at the surface of FeS(s) minerals, and the two crystalline forms cinnabar and metacinnabar, α-HgS(s) and βHgS(s), respectively.17 β-HgS(s) can form in situ in sediments and soils and α-HgS(s) may be present as a consequence of transport and deposition of weathered minerals. The speciation of adsorbed and solid phase HgII is potentially a key factor Received: Revised: Accepted: Published: 11653

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tracers (separate experiment) were not statistically different (ANOVA, p > 0.05). Tracer slurries were added to 20−50 g of sediment (n = 3) under an N2 atmosphere using a glovebox, which gave the final concentrations shown in Tables S1 and S2. The sediment was manually shaken and split into subsamples (3−5 g each) and incubated at room temperature in the dark under an N2 atmosphere for varying time periods. Incubation was terminated by sample freezing. The amount of MeHg from added HgII tracers in the T0 samples (with incubation time “0 h”) were always below or close to the detection limit, which shows that possible methylation during the time required to freeze the samples was negligible. The absolute rates determined in sediment slurries in the laboratory may deviate from the rates in intact cores under natural conditions. Our results (Results and Discussion) do however clearly demonstrate that the use of geochemically relevant tracers provides reaction rates close to those for ambient Hg in the experimental system, and that these rates can deviate up to 2 orders of magnitude compared to rates determined from traditionally used dissolved tracers. Me204Hg was added as an internal standard to samples and blanks (MQ water) and left for 1 h for equilibration prior to extraction. The MeHg was then double extracted (CuSO4/ KBr/H2SO4 followed by CH2Cl2) into MQ water, derivatized with NaB(C2H5)4, purged and trapped on Tenax TA sorbent.20 The samples were then analyzed by thermal desorption gas chromatography−inductively coupled plasma mass spectrometry.26 Concentrations of ambient MeHg and MeHg formed from added HgII isotope tracers were calculated from mass-biascorrected signals using a set of linear equations described by Qvarnström et al.27 The HgII methylation rate constant and the MeHg demethylation rate constant were calculated by fitting the concentration of MeHg formed from each tracer during the first 7 d of incubation with a nonlinear reversible reaction model (eq S3, SI text). In contrast to a more commonly applied approach based on eq S2, the approach applied here allows accurate determination of km even if both formation and degradation of MeHg, as well as variations in the HgII and MeHg concentrations, are significant during the incubation experiment. The amount of formed MeHg was compared for filtered (0.02 μm, Anatop10, Whatman) (n = 3) and unfiltered Hg-X(s) slurries (n = 3) at concentrations that corresponded to 0.07−0.2 times the ambient Hg concentration. Separate sediment subsamples were incubated with only HgII(aq) (n = 3) or HgII(s) (n = 3) tracers, respectively, for up to 2 d to confirm that the use of multiple isotope tracers did not cause any interfering effects on the results. Further, incubations with higher concentrations (0.4−3 and 9−14 times ambient Hg) of solid tracers were evaluated to optimize the concentration level of the tracers. Additional experiments to validate the proposed methodology are described in SI text.

controlling MeHg formation rates, but knowledge about quantitative differences in methylation rate between different adsorbed/solid HgII phases in natural systems is lacking. Isotope tracer methodologies are one of the most powerful experimental tools for the study of MeHg formation and degradation and the use of such methods has, over the past decade, significantly increased the understanding of these processes.10,18−20 For sediments and soils the experimental conditions are however intricate because HgII is mainly present as the above listed solid phases or surface complexes controlled by its strong affinity for reduced organic and inorganic sulfur groups. In contrast, isotope tracers have traditionally been introduced as labile aqueous complexes.21 Information of how well aqueous tracers yield realistic reaction rates for ambient HgII and MeHg is scant. Strict experiments on the kinetics of Hg precipitation− dissolution and adsorption−desorption with major inorganic and organic solid and surface complexes in soils and sediments in which the experiments are not operationally linked to a certain method (e.g., Lamborg et al.)22 are very few.23 It has, however, generally been assumed,16,24 and in some studies experimentally demonstrated,15,19,25,26 that labile aqueousphase isotope tracers react at higher rates than ambient HgII. As a consequence, the terms potential, specif ic, or conditional have been used to acknowledge the expected overestimated rates and rate constants determined from added isotope tracers.21 Despite this limitation, added tracers have been shown to exhibit temporal trends in reactivity and treatment effects similar to those of ambient Hg in incubation experiments, even if the absolute rates deviate.15,19,26 In this study, we propose a new methodology in which major adsorbed and solid phases of HgII in sediments are utilized as isotope tracers. With this methodology we determined (1) species-specific HgII methylation rates for geochemical forms in control of aqueous phase concentrations of HgII in sediments, and, based on species-specific rates, (2) a methylation rate relevant for ambient HgII in an estuarine sediment.

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MATERIALS AND METHODS Sediments (0−5 cm) were collected at the estuary of Ö re river in the Bothnian Bay (coordinates: 63° 33.905′ N, 19° 50.898′ E) at a water depth of 13 m using a core sampler. Samples were collected once and stored in the dark at 4 °C. Separate sample batches (denoted batches I, II, III, and IV in Tables S3 and S4 in the Supporting Information (SI)), with n = 3 for each incubation time and treatment, were withdrawn for incubation experiments. No systematic (e.g., p = 0.3 for the 198Hg(NO3)2(aq) tracer) effect on determined km values was observed during 6 months of experiments. The following isotopically enriched HgII tracers were prepared in-house (SI text): α-199HgS(s), β-201HgS(s), HgII reacted with Mackinawite (≡FeS-200HgII or ≡FeS-202HgII), HgII bound to natural organic matter (NOM-196HgII), and the traditionally used aqueous HgII nitrate solution, 198Hg(NO3)2(aq). Detailed descriptions of the characterization of synthesized solid-phase tracers using XRD, XPS, SEM, and EXAFS are given in the SI. Briefly, XRD measurements indicated pure α-HgS(s) and β-HgS(s) crystalline structures for these tracers (SI text, Figures S1, S2), whereas XPS analysis (SI text) indicated two additional structures other than HgS at the α-HgS(s) surface (Table S1, Figure S4). These additional structures were removed when the precipitate was washed with 7 M nitric acid (Figure S6), and the km values obtained with washed and unwashed α-199HgS(s)

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RESULTS AND DISCUSSION Methylation Rate for Different HgII Tracers. Solid and aqueous HgII tracers were added to estuarine sediments at concentrations that were 0.039−0.24 (Figure 1, Table S3) or 0.14−2.6 (Table S4) times the ambient total Hg concentration (104 ng g−1 dry mass). The highest methylation rate was obtained for the traditionally used aqueous tracer 198Hg(NO 3 ) 2 (aq), intermediate rates for NOM- 196 Hg II and ≡FeS-202HgII, and the lowest rates were obtained for the two solid-phase HgS(s) tracers (Figure 1). The HgII methylation 11654

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Table 1. Average HgII Methylation Rate Constants (± SE) (from 4 Batches with n = 3 per Batch), km (d−1), Determined for Each Isotope Tracer Added to an Estuarine Sediment Sample at Four Different Occasions at Concentrations That Corresponded to 0.039−2.6 Times the Ambient Total Hg Concentration Hg tracer 198

Hg(NO3)2(aq) NOM-196HgII ≡FeS-202HgII α-199HgS(s) β-201HgS(s) ambient HgII-calcc

methylation rate constant ± SEa, km (d−1) 0.12 0.029 0.014 0.0066 0.0012 0.010

± ± ± ± ± ±

0.014 ab 0.0030 b 0.0033 c 0.0013 c 0.0013 d 0.0011

a

For details on the individual methylation rate constants at each incubation occasion, see Table S3. bDifferent letters indicate statistical significant difference (p < 0.05) according to a Post-Hoc pair wise t test following one-way ANOVA with incubation occasion (n = 4) as dummy variable. Due to skewed distribution the statistical test was conducted on logarithmated data. ANOVA statistics: n = 60; r2 = 0.89; r2-adj = 0.84; RMSE = 0.72; p < 0.001; model SS = 164; error SS = 21; SS total = 188; F-ratio = 16.8. cThe km of ambient HgII-calc was calculated from the rate constant of a mixture of tracers that corresponded to the solid-phase speciation of ambient HgII in the sediment as determined by HgII-EXAFS (30% NOM-HgII and 70% βHgS(s)).

Figure 1. Molar fraction (±SE, n = 3) of MeHg formed as a function of time from each HgII tracer added to an estuarine sediment: (a) 198 Hg(NO3)2(aq) (□), NOM-196HgII (●), ≡FeS-202HgII (▲), 199 α- HgS(s) (⧫), and β-201HgS(s) (■); (b) close-up of α-199HgS(s) and β-201HgS(s).

rate constant, km, was calculated by fitting a nonlinear, reversible reaction model (eq S3) to the concentrations of MeHg formed from each tracer during the first 7 days of incubation (c.f. Hintelmann et al.,17 Martin-Doimeadios et al.19,28). The calculation approach used allows the accurate determination of km, even if both the formation and degradation of MeHg, as well as variation in the HgII concentration, are significant during the incubation experiment (SI text). The use of 198Hg(NO3)2(aq) yielded an average km value (Table 1, 4 batches; n = 3 per batch) of 0.12 ± 0.014 (SE) d−1, which was 4−9 and ∼20−100 times greater (p < 0.5) than the km determined for NOM-196HgII and ≡FeS-202HgII, and α-199HgS(s) and β-201HgS(s), respectively. The inter- and intrabatch reproducibility for determined km values were in general good for all tracers except for ≡FeS-HgII, see Figure 2. Freshly prepared ≡FeS-202HgII exhibited an availability for methylation similar (p > 0.05) to that of thiolbound HgII in NOM, whereas ≡FeS-202HgII aged for 24 days (after addition of 202HgII(aq) to the FeS(s)) exhibited an availability similar to that of β-HgS(s) (Table S3). Adsorption29 and Hg EXAFS30 studies have shown that HgII(aq) may either form surface complexes with the FeS(s) phase or dissolve it through the formation of β-HgS(s), depending on the aging time of the FeS(s) phase prior to the addition of HgII. Freshly prepared (within a few hours) FeS(s) is highly distorted and will be readily dissolved by HgII, especially in the presence of NOM.30 With aging (days), the FeS(s) becomes more

Figure 2. Average methylation rate constant (±SE, n = 3), km (d−1) determined for each isotope tracer added to four separate batches of an estuarine sediment sample at concentrations that corresponded to 0.039−0.24 (bars no. 1−3 from left for each tracer) or 0.14−2.6 (bar no. 4 from left for each tracer) times the ambient total Hg concentration.

crystalline and thus chemically more stable, which may lead to the formation of ≡FeS-HgII surface complexes at low HgII surface coverage.29 The FeS(s) used in this study was aged for 10 or 14 days prior to the addition of HgII(aq), which suggests conditions that favor surface complexation. It is reasonable to assume (although this has not been explicitly studied), however, that aged FeS(s) may also be gradually dissolved by HgII with time through the formation of β-HgS(s). This dissolution would explain the lower (close to significant, ANOVA, p = 0.056) methylation rate for the “aged” (24 days) compared to the “fresh” ( 0.05) different km values but with a lower relative uncertainty for the higher additions (Table S4). Higher concentrations (at least up to two 11657

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to establish if a NOM-HgII or a β-HgS(s) tracer, or a mixture of both of these, should be used. This can be predicted from chemical equilibrium modeling (e.g., SI text) based on measurements of total Hg concentrations and ancillary chemical parameters likely to be included in monitoring studies. The most critical issue of this modeling approach is the selection of relevant stability constants, which should be done carefully.16

times the total ambient Hg concentration, depending on the tracer used) than those typically used in traditional tracer studies can thus be applied. Higher concentrations are recommended (to obtain low relative uncertainties) for solidphase tracers because the bioavailable fraction of these tracers is much smaller than that of a tracer in aqueous form. The highest concentration range of solid tracers inhibited MeHg formation during the first 2 days of incubation, which prevented the calculation of km values. However, after 2 days, MeHg formation did progress, and the amount formed after 7 days was included as the third data point in Figure 4. MeHg formation in the sediment during 7 days of incubation with the lowest tracer concentrations was compared for HgII solid-phase tracer slurries that were filtered (0.02 μm) or unfiltered prior to addition of the suspensions to the sediment samples. The use of filtered tracer slurries suppressed MeHg formation compared to unfiltered tracers by 88% for NOM- 196 Hg II , 70% for α- 199 HgS(s), 100% for both ≡FeS-200HgII and β-201HgS(s), and 12% for 198Hg(NO3)2(aq). This shows that for all solid and adsorbed tracers, most or all MeHg was formed from HgII substrate originally added to the sediment sample as solid or adsorbed phases with particle size larger than 0.02 μm. Interestingly, low methylation was observed for the filtered α-199HgS(s) tracer slurry, which implied that part of the Me199Hg formed from this tracer originated from dissolution of HgII (during incubation) from nanoparticles that passed through the 0.02-μm filter. Assuming a higher solubility/dissolution rate for nano compared to microparticulate HgS(s), as discussed above, this might explain the higher (p < 0.05) km values obtained for α-HgS(s) compared to β-HgS(s) in this study. The possibility that the Me199Hg formed from the filtered α-199HgS(s) originated from HgII dissolved in the tracer slurry prior to incubation could be excluded because the amount of Me199Hg formed (0.013 ng g−1 dry mass) exceeded the calculated (solubility product, Ksp = 2 · 10−32) amount of dissolved HgII in equilibrium with α-HgS(s) by ∼1018 times. To verify that the addition of a solid and aqueous tracer mixture did not cause interference in the determination of km, incubation experiments were also conducted with additions of 198 Hg(NO3)2(aq) and solid-phase tracers to separate sediment subsamples. The differences in km determined in the single and multitracer experiments were not statistically different for the solid HgII tracers (ANOVA, p > 0.05). For the 198Hg(NO3 ) 2 (aq) tracer, a small but statistically significant (ANOVA, p < 0.05) deviation of ∼10% was observed. These results showed that the proposed methodology was robust and reproducible and clearly demonstrated methylation of HgII substrate that originated from truly adsorbed and solid phases. The proposed methodology is not only useful for refined studies on MeHg formation from different specific solid HgII phases, but can also replace the traditional tracer approach (using labile aqueous tracers) in large-scale survey and monitoring studies on MeHg formation, as discussed above. To fully mimic the geochemical conditions in a particular environment by the solid-phase tracer methodology it is necessary to have knowledge, or be able to make qualified assumptions, about the solid-phase speciation of ambient HgII in the samples. This is not trivial since available approaches (e.g., Hg-EXAFS measurements, sequential extraction methods, or chemical equilibrium modeling) all have constraints. However, in most sample types relevant for HgII methylation (suboxic sediments and soils) the issue will likely be restricted

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ASSOCIATED CONTENT

S Supporting Information *

Detailed descriptions of (i) preparation of solid-phase isotope tracers, (ii) XRD, XPS, and SEM characterization of synthesized HgS(s) tracers, (iii) calculation of the HgII methylation rate constant, km, (iv) Hg LIII-edge EXAFS determination, and (v) solid-phase chemical speciation calculation of HgII. This information is available free of charge via the Internet at http://pubs.acs.org/

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AUTHOR INFORMATION

Corresponding Author

*Phone: +46 90 786 5189; e-mail: [email protected]. Notes

The authors declare no competing financial interest.

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ACKNOWLEDGMENTS This work was supported by the Swedish Research Council (grant 2008-4363), Umeå Marine Sciences Centre, Umeå University, the Kempe Foundation, and the Knut and Alice Wallenberg Foundation. We thank Dan Boström and Per Hörstedt, Umeå University, for XRD and SEM measurements, respectively.

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