Mechanisms of methyl mercury net degradation in alder swamps: the

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Mechanisms of methyl mercury net degradation in alder swamps: the role of methanogens and abiotic processes Rose-Marie Kronberg, Jeffra K Schaefer, Erik Björn, and Ulf Skyllberg Environ. Sci. Technol. Lett., Just Accepted Manuscript • DOI: 10.1021/acs.estlett.8b00081 • Publication Date (Web): 14 Mar 2018 Downloaded from http://pubs.acs.org on March 16, 2018

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Environmental Science & Technology Letters

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Mechanisms of Methyl Mercury Net Degradation in Alder Swamps: the Role of Methanogens and Abiotic Processes

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Rose-Marie Kronberg1, Jeffra K. Schaefer2, Erik Björn3, and, Ulf Skyllberg1*

8 1Department

of Forest Ecology and Management, Swedish University of Agricultural

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Sciences, SE-901 83 Umeå, Sweden

12

2Department

of Environmental Sciences, Rutgers University, 14 College Farm Road, New

Brunswick, NJ 08901

14 3Chemistry

Department, Umeå University, SE-901 87 Umeå, Sweden

16 *Corresponding author: Ulf Skyllberg. Phone: +46 (0)90-786 84 60; e-mail:

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[email protected]

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1 ACS Paragon Plus Environment


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ABSTRACT Wetlands are common net producers of the neurotoxin monomethylmercury (MeHg)

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and are largely responsible for MeHg bioaccumulation in aquatic food-webs. However, not all wetlands net produce MeHg - notable exceptions are black alder (Alnus glutinosa)

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swamps which net degrade MeHg. Here we report the mechanisms of MeHg demethylation in one such swamp (EHT), shown to be a sink for MeHg during four

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consecutive years. The potential demethylation rate constant (kd) in soil incubations was ~3 times higher in the downstream (EHT-D: kd ~0.14 d-1) as compared to the

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upstream part of the swamp (EHT-U: kd ~0.05 d-1). This difference concurred with increased stream and soil pH, and a change in plant community composition. Electron

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acceptor and inhibitor addition experiments revealed that abiotic demethylation dominated at EHT-U while an additional and equally large contribution from biotic

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degradation was observed at EHT-D, explaining the increase in MeHg degradation. Biotic demethylation (EHT-D) was primarily due to methanogens, inferred by a decrease in kd

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to autoclaved levels following selective inhibition of methanogens, not sulfate-reducers. While methanogen-specific transcripts (mcrA) were found throughout the wetland,

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transcripts clustering with Methanosaetaceae were exclusive to EHT-D, suggesting a possible role for these acetoclastic methanogens in the degradation of MeHg.

40 INTRODUCTION

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Mercury (Hg) in general, but more specifically the neurotoxin monomethylmercury (MeHg), tends to bioaccumulate and biomagnify in the aquatic environment. Exposure

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from fish consumption is of particular concern to humans, and in Sweden alone Hg concentrations in fish are above consumption guidelines (0.02 mg kg-1 ww)1 in most of

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its ~100 000 lakes, and the problem is persisting.2,3 The amount of MeHg available for bioaccumulation is the net of two concurrent processes; methylation of inorganic

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divalent Hg (HgII) and MeHg demethylation. Methylation is primarily an anaerobic


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process while demethylation occurs in both anaerobic and aerated environments.4–6 Our

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understanding of biotic methylation mechanisms has greatly improved over the last years, including identification of the two genes hgcA and hgcB by which MeHg is formed

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in bacteria and archaea.7 Our understanding of mechanisms and factors in control of MeHg degrading processes in natural environments are however less developed.

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Both abiotic and biotic demethylation occur, with photolysis of MeHg in surface waters being the major sink of MeHg in aquatic environments,8 and the focus of most

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MeHg degradation studies.9 There are two known pathways for biotic demethylation; 1) the combined actions of the organomercurial lyase (MerB) and mercuric reductase

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(MerA),10 converting MeHg to Hg0 and CH4, and 2) an “oxidative” pathway where MeHg is degraded to HgII and either CO2 and/or CH4, allowing for potential recycling of HgII

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back to MeHg.4 The former pathway dominates in Hg contaminated areas while the latter is more common at unpolluted sites.5,6 Aerobic organisms known to degrade MeHg

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via the mer pathway10 have recently been shown to degrade MeHg independent of mer, including some methanotrophs11 and a Pseudomonas sp.12 In anoxic wetlands and

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sediments, both methanogens and sulfate-reducing bacteria (SRB) have been implicated in the oxidative demethylation of MeHg4,13 and some SRB strains with this ability have

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been identified.14 No methanogen capable of demethylation is known and only one isolate has been tested.15 Iron reducing bacteria (IRB) have been inferred as

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demethylators in sediment samples16 and one strain, Geobacter bemidjiensis was recently shown to demethylate MeHg in a laboratory experiment.17

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Anaerobic stimulation of HgII methylation dictate wetlands as sources of MeHg to downstream environments.18,19 However, a black alder (Alnus glutinosa) swamp

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(Edshult: EHT) was established as a significant and consistent MeHg sink during four years,19 and a follow-up synoptic screening implied Alnus swamps acting as MeHg sinks

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in general.20 Further, the studies at EHT demonstrated decreasing MeHg pools alongside



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increasing potential demethylation rate constants in soils along the water pathway

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through the wetland. Here we investigated the mechanisms of MeHg demethylation in the soils of

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EHT, using experimental incubations with amendments of electron acceptors and specific microbial inhibitors, and an autoclaved control to determine the contribution

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from abiotic processes alone. We also constructed clonal libraries and used gene transcription data, attempting to reveal the microbial communities degrading MeHg.

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Our results clarify that abiotic demethylation is widely distributed with an equal contribution from biological demethylation in the downstream part of the swamp where

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rates of demethylation peaked. Methanogen transcripts were abundant and demethylation was strongly inhibited by bromoethanesulfonic acid (BES); thus implying

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methanogens as important for the biological demethylation. Because of the ubiquity and activity of methanogens in soils and wetlands, our findings provide an important leap

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forward towards a mechanistic understanding of the MeHg dynamics on the landscape level. Similar to wetlands serving as nitrogen and phosphorous traps in order to protect

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surface waters from eutrophication,21 wetlands with the ability to degrade MeHg20 should be recognized and actively used for mitigation purposes.

92 MATERIALS AND METHODS

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A short description of methodology follows with more details in the Supporting Information, where also the study sites are described briefly. All sites have been

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described in detail elsewhere.20,22 Soil sampling and handling. Soil and stream waters were sampled at all sites (seven

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Alnus swamps and two Sphagnum peatlands) in May 2010 and at the Alnus swamp EHT in August 2012, following clean sampling protocols. Samples were double bagged, then

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put in a cooler on ice and transported to the lab the same day (2012) or within a week (2010). In the lab, samples were refrigerated at 4°C until transferred to a N2 filled


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glovebox (COY) at least 24h prior to incubation experiments. Ancillary chemistry, and MeHg and total Hg (HgTOT) analyses are described previously.20,22

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Incubation studies. Enriched stable Hg isotope tracers were used for the determination of potential methylation and demethylation rate constants (km and kd), following a

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previously described protocol.23,24 Briefly, soil was homogenized in the glovebox and distributed (~10 g) to 50 mL centrifuge tubes (Sarstedt). Incubations in 2012 were

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prepared in triplicate for each treatment. For monitoring demethylation, isotope tracers (Me198HgCl (92.78%) and Me204HgCl (91.67%) in 2010 and 2012, respectively) were

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mixed into the soil and then split into two tubes; T0 and T48. The T0 tubes were immediately put on dry ice and stored frozen at -20°C until analysis. The T48 tubes were

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incubated dark, at room temperature for 48 h after which they were frozen at -20°C. Subsamples (1.0 g) were collected at T0 and T48 for molecular analysis. Details

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regarding the kd calculations, incubation studies, RNA/DNA extraction, RT-PCR amplification, cloning, sequencing, and phylogenetic classification are provided in

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Supporting Information. Sample incubations in 2012 were supplemented with SO4, colloidal FeIII, and

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colloidal MnIV as electron acceptors. We used MoO4 and BES as specific inhibitors of SRBs and methanogens respectively.25 Abiotic demethylation was examined using

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autoclaved soil. Ambient levels of sulfate was ~20µM and Fe was ~0.6 mM, based on empirical data from previous years.20,22 Sulfate was added at a final porewater

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concentration of 20 times ambient, and molybdate was added at the same concentration to keep the total amendment low and thus specific to SRB. Amounts of added FeIII and

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MnIV corresponded to 2 mmol/L porewater and were added as a suspension in MQ. The BES addition had a final concentration of 10 mM in porewater.

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Statistics. The data was analyzed using the software Graphpad Prism (USA). For the incubation studies, ANOVA was used to test for significant differences between control

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and treatments. If the null hypothesis was rejected, analysis proceeded with Dunnett’s



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multiple comparisons test, with 7 comparisons in total. For the regression analyses, the

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data was log-transformed (natural logarithm) for normality before significance analysis. All tests were carried out at the 0.05 significance level, and we used the significance

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level of 0.10 as marginally significant.

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RESULTS AND DISCUSSION Demethylation in Alnus swamps in contrast to Sphagnum peatlands. The Alnus

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swamps in this study have previously been indicated as net MeHg sinks,19,20 and site Edshult (EHT) has proved to be a consistent net MeHg sink during four consecutive

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years. Here we determined kd in anoxic soils sampled in the upper, lower and western parts of Edshult (EHT-U, EHT-D, and EHT-W), as well as in the center of six additional

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Alnus swamps (Kolsboda, Kvillehult, Klasentorp, Löneberga, Steglehylte, and Nybygget), see Figure S1. Demethylation rates were also determined in the bog and fen parts of the

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Sphagnum peatland Långedalen (LDN-B and LDN-F), and in one other Sphagnum peatland (Ystebo). All sites are located across southern Sweden.

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Molybdate was added to incubations to specifically inhibit SRB, as previous studies with freshwater sediments,4,13,26 periphyton,27 and in culture14 suggest SRB

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involvement in MeHg degradation (Figure 1). However, with only two exceptions (EHTW and Steglehylte), molybdate not only failed to inhibit demethylation but had the

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opposite effect; stimulating MeHg degradation at the Alnus sites. In contrast, at the peatland sites molybdate inhibited demethylation at four sites out of five. The lack of

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inhibition by molybdate at the Alnus sites agrees with our earlier findings at EHT,22 and indicates that in the native soil, SRB are competing with the MeHg demethylating

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community for energy and/or nutrients. Since methanogens often are stimulated by molybdate addition, due to shared electron donors (e.g. H2, formate)25,27–29 and similar

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energy yields, these results imply methanogens in the demethylation of MeHg, as reported previously.4,13,26 The inhibition of demethylation by molybdate in the less


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productive peatlands may be indicative of either direct demethylation by SRB or inhibition of methanogens metabolizing syntrophically with SRBs.4,30

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Regression analyses (post hoc) suggest that kd was positively related to pH, within the interval 4.7 – 6.4 (r2=0.63, p=0.018), and negatively to the C/N (r2=0.39,

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p=0.097, marginally significant), Figure S2. Although these relationships are based on relatively few data points, they agree with the conceptual model for boreal wetlands

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proposed by Tjerngren et al.19,22 According to this model, the km/kd ratio shows an optimum at intermediate trophic status (reflected by C/N) and acidity (reflected by pH),

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where the decrease is mainly driven by increased demethylation as enhanced with improved nutrient conditions. With the exception of sites Klasentorp and Löneberg our

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wetlands followed this model well (Figure S3). Spatial variability in demethylation mechanisms across the Alnus swamp Edshult.

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Triplicate anoxic soil incubations showed that the control sample kd in the downstream part of Edshult (EHT-D) was two to three times higher than in the upstream portion

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(EHT-U) near the inflow (~0.14 d-1 and ~0.05 d-1, respectively: Figure 2), well in agreement with our previous studies at this site where the same pattern was observed

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in 2006, 2008, and 2009.20 Further, the abiotic (autoclaved) demethylation was similar at EHT-D and EHT-U (0.06 – 0.08 d-1). We cannot exclude the possibility that autoclaving

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caused physiochemical changes, somehow enhancing demethylation and masking biological demethylation at EHT-U. However, the remarkable similarity in kd between

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the autoclaved incubations (across the wetland), most live incubations at EHT-U, and BES-inhibited incubations at EHT-D are not likely coincidental. Instead, a simpler

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explanation is that abiotic MeHg degradation is a common process throughout the swamp with biological demethylation adding to the increased demethylation at EHT-D.

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Potentially stimulatory (SO4, FeIII, and MnIV) or inhibitory (MoO4 or BES) amendments were also investigated (Figure 2). Similar to the 2010 incubations

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discussed above (Figure 1), molybdate had either no effect (EHT-U) or slightly



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stimulated (EHT-D) demethylation. The addition of BES inhibited demethylation at EHT-

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D and decreased kd down to autoclaved levels (~0.08 d-1), providing strong support for methanogens degrading MeHg here. Similar findings were reported in sediments of the

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Florida Everglades.4,5 At EHT-D, FeIII gave significantly higher kd compared to the control (p

Mechanisms of methyl mercury net degradation in alder swamps: the

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