roduction in a chronically sulfate-impacted sub-boreal wetland

Increased deposition of atmospheric sulfate exacerbates methylmercury (MeHg) production in freshwater wetlands by stimulating methylating bacteria, but it is unclear how methylation in sub-boreal wetlands is impacted by chronically elevated sulfate inputs, such as through mine discharges. The purpose of our study is to determine how sulfate discharges to wetlands from iron mining activities impact MeHg production. In this study, we compare spatial and temporal patterns in MeHg and associated geochemistry in two wetlands receiving contrasting loads of sulfate. Two orders of magnitude less sulfate in the un-impacted wetland create significant differences in acid-volatile sulfide and porewater sulfide; however, dissolved and solid-phase MeHg concentrations and methylation rate potentials (Kmeth) are statistically similar in both wetlands. Permitted mine pumping events flood the sulfate-impacted wetland with very high sulfate waters during the fall. In contrast to observations in sulfate-limited systems, this large input of sulfate to a chronically sulfate-impacted system led to significantly lower potential relative methylation rates, suggesting a predominance of demethylation processes over methylation processes during the sulfate loading. Overall, short-term measurements of methylation and demethylation potential are unrelated to gross measures of long-term MeHg accumulation, indicating a decoupling of shortand long-term process measurements and an overall disequilibrium in the systems. High sulfide accumulation, above 600–800 mg l 1 sulfide, in the sulfate-impacted system lowers long-term MeHg accumulation, perhaps as a result of less bioavailable Hg–S complexes. Although continued research is required to determine how sulfate-limited freshwater wetlands might respond to new, large inputs of high-sulfate runoff from mining operations, chronically impacted wetlands do not appear to continually accumulate or produce MeHg at rates different from wetlands unimpacted by mining.


Introduction
The bioaccumulation of mercury (Hg) in aquatic ecosystemsa serious global concern owing to its toxic effects in humans and wildlife 1,2 has led to sh consumption advisories for thousands of lakes across the US 3,4 and recent worldwide efforts to curb Hg emissions and exposure through the Minimata Convention. 5 Methylmercury (MeHg) is the bioaccumulative form of Hg, and its production in aquatic ecosystems is a key factor controlling the organismal load of Hg. 6 A signicant amount of research therefore has been devoted to understanding the complex set of interrelated chemical and biological processes that produce, break down, and transport MeHg in a variety of natural settings. 7,8The production of MeHg in natural environments is primarily an anaerobic, microbially-driven process. 9lthough a wide array of anaerobes harbor the two genes (hgcAB) responsible for Hg methylation capabilities, 10 sulfate-reducing bacteria (SRB) have consistently dominated other functional groups as conclusive Hg methylators. 11,12etlands, with their highly organic soils and saturated hydrologic conditions, harbor a myriad of anaerobic microbes and are important net producers of MeHg. 13Considerable output of MeHg to downstream ecosystems can occur when areas of relatively high MeHg production and accumulation are linked to hydrologically connected areas. 14,15In freshwater wetlands, MeHg production appears limited by sulfate availability, particularly in nutrient poor peatlands, 16,17 and Hg methylation capabilities vary considerably across different wetland types. 18,19Discrete spatial variations in sulfate loads and redox processes within wetlands are important to the internal spatial patterning of MeHg production. 15Thus, the production of MeHg in wetlands is important in dening in situ wetland MeHg concentrations and also as a source for downstream rivers and lakes.
The role of sulfate loading on MeHg production in wetlands with chronically elevated sulfate concentrations is much less clear than for sulfate-limited wetlands.For example, in the Florida Everglades where sulfate is a common agricultural fertilizer, sediment exposed to a range of sulfate and nutrient loads showed lower net MeHg production in areas where sulfate and nutrient loads were already elevated. 20In the Prairie Pothole Region of Saskatchewan, where freshwater sulfate concentrations can be naturally very high (up to 1000 s of mg l À1 ), Hoggarth et al. 21found no signicant difference in Hg methylation potentials between high-sulfate (>150 mg l À1 ) and low-sulfate (<20 mg l À1 ) wetland sediment.The buildup of sulde in anoxic sediment and porewater exposed to high sulfur loads is thought to alter the form of inorganic mercury, reducing its bioavailability to methylating microbes due to either thermodynamically or kinetically limited processes. 7,22ron ore mining has been an important economic activity on the Mesabi Iron Range, in the St. Louis River watershed of northern Minnesota since the early 20 th century, 23 with current activities focused on mining and beneciation of taconite, a low-grade ore.Over the years, there has been a signicant accumulation of sulfate in tailings ponds (constructed settling basins for ne-grained waste rock) and pit lakes (abandoned open-pit mines) resulting from a mixture of sources including ore and waste rock sulde mineralization 24 and the taconite pellet induration process.25 Non-point discharges, including water leaking from tailings basins and runoff from tailings piles, as well as permitted discharges from active dewatering operations and pit-lake overows have signicantly increased sulfate loading to the St. Louis River, which drains into western Lake Superior.26,27 While abundant carbonate minerals in the iron formation appear to fully neutralize acid generation, sulfate concentrations above 100 mg l À1 have been observed in local streams and in excess of 1000 mg l À1 in some mine pits.27 Background surface water sulfate concentrations in the region are less than 5 mg l À1 .Despite receiving far greater loads of sulfate, tributary streams impacted by mining in the St. Louis River watershed carry similar concentrations of MeHg as their non-sulfate impacted counterparts 27 and do not signicantly differ with respect to Hg bioaccumulation by dragony larvae.28 Our study was designed to investigate how elevated sulfate loading from mining affects MeHg production in an impacted wetland and to compare with Hg methylation capabilities in a wetland hydrologically disconnected from elevated sulfate loading.2

Site description
This study focused on the St. Louis River watershed, which drains through northeastern Minnesota into western Lake Superior (Fig. 1).The regional climate is humid continental, with an average annual temperature of 3.5 C and average annual precipitation of 680 mm.Two wetland sites with contrasting upstream mining impacts were selected for the study.The Long Lake Creek (LLC) wetland (N 47.42 , W 92.56 ) is a $15 hectare sub-boreal fen that receives sulfate-rich water (typically $250 mg l À1 ) from a $2.5 km 2 , mining-inuenced catchment.Previous work indicates signicant sulfate reduction as waters pass through the wetland. 29The LLC periphery supports typical peatland vegetation with Sphagnum dominated peat, grading into a degraded fen margin dominated by Typha spp., and a 1.5 hectare open-water pool.The un-impacted West Two River (WTR) wetland (N 47.465 , W 92.77 ) lies near the headwaters of a small (1.2 km 2 ) catchment with no ongoing mining operations and drains into the west branch of the West Two River.WTR is a large sedge-dominated fen on the north side of a small (4.5 hectare) pond with surface water sulfate concentrations typically less than 5 mg l À1 .
The study wetlands are similar in size and in that both surround open water pools with small, but active riverine inputs and outlets.Our study year (2012) was characterized by several large rainfall events in June ($300 mm) and mostly dry conditions through the remainder of the summer (<150 mm until October).In both wetlands, the water table ranged from 10-20 cm above the peat surface in late June and early July to approximately 5-10 cm below the peat surface in September.Following this dry period, the LLC wetland was subject to a ooding event from permitted pumping and discharge of water from an upstream mine pit in mid-September.In accordance with its discharge permit, the mining company reported that the ow in the Long Lake Creek watershed increased from near zero to 0.25-0.38m 3 s À1 during the week following September 15, 2012. 30The pumping lasted through November and resulted in an inundation of the entire wetland area with 30-50 cm of sulfate-rich (>500 mg l À1 ) water.This same type of pumping event occurred during 2011 and also possibly in years prior, meaning the LLC wetland complex was inundated by high sulfate water during the fall preceding the data reported here for 2012.

Sampling methods
Samples were collected from two sampling plots at each wetland site: (1) from an area 5-10 m adjacent to the open water pool (hereaer "near open water"), and (2) within approximately 5 m of the surrounding terrestrial upland (hereaer "near upland," Fig. 1c).Peat samples were collected in May, July, and October 2012 from the Long Lake Creek wetland (hereaer "sulfate-impacted") and in July and October 2012 at the West Two River wetland (hereaer "sulfate-unimpacted").There are no samples from the sulfate-unimpacted site during May because we originally chose a different site that was not well-matched geomorphically to the LLC wetland and subsequently abandoned the site for West Two River starting in July.
For methylation and demethylation assays, triplicate 5 cm diameter cores were collected at each plot using clear polycarbonate tubes, capped, and stored upright in coolers during transport to a local lab.Core tubes were specially designed with silicone-sealed injection ports spaced at 1 cm intervals to allow for enriched Hg isotope injections (explained further below).For other solid phase analyses, triplicate peat blocks were cut with a serrated knife from each sampling plot, wrapped in plastic wrap to preserve moisture, and transported to a local lab for sectioning and processing.The center portion of moist peat blocks were sliced into 0-2, 2-4, and 4-8 cm depth sections and immediately placed into an oxygen-free atmosphere where individual samples were homogenized.Sediment samples for acid volatile sulde (AVS) analysis were sealed in a nitrogen atmosphere, preserved with zinc acetate (nal concentration of 0.02 mol l À1 ), and frozen until analysis.
Porewater samples were collected in triplicate from each sampling plot in May 2012 and then approximately biweekly during July, August, and September using 10 cm Rhizon® lters 31 installed in situ and connected to acid-washed, evacuated serum bottles with PTFE tubing and stainless steel hypodermic needles.
In the case of a water table at or above the peat surface, porewater was collected from the surcial 10 cm of peat pore uids.When the water table was below the peat surface, porewater was collected from the uppermost 10 cm of saturated peat.Replicate Rhizons® were typically installed in the aernoon and retrieved the following morning, allowing time for the in situ ltration of four bottles totaling approximately 325 ml.One bottle was pre-loaded with ZnAc for the preservation of sulde (nal concentration 0.002 mol l À1 ), while other bottles containing no preservative were opened in the lab, allocated, and preserved as needed for analysis of total mercury (THg), MeHg, and sulfate.In an effort to avoid sampling of disturbed areas, an approximately 7 meter diameter area was staked out at each plot at the beginning of the season and samples were collected from different areas within the plot during each sampling event.Because previous research has shown that microtopographic characteristics of peatlands can have signicant inuence on mercury and related geochemistry, 32 an effort was made to consistently collect both solid and porewater samples from representative hollows between hummocks that appeared similarly moist or inundated compared to other hollows in the immediate area.Total-and methyl-mercury samples were preserved with trace metal grade 12 N HCl (0.5% by volume) and stored on ice in double-bagged PETG bottles.Porewater samples for sulfate analysis were acidied to pH <3 with concentrated HCl and gently bubbled with oxygen-free nitrogen to remove dissolved sulde prior to storage at 4 C until analysis.

Analytical methods
Methylation/demethylation potentials and mercury analyses.Mercury methylation and demethylation rate potentials were assessed using enriched stable mercury isotope injection and incubation techniques as outlined in Hintelmann et al. 33 and Mitchell and Gilmour. 18Sediment cores were injected with a mixture of stable isotope-enriched 200 Hg 2+ and Me 201 Hg + (94.3% 200 Hg 2+ and 84.7% Me 201 Hg + ) using Hg-Cl ligand stock solutions that were equilibrated for one hour with ltered site porewater that had not been exposed to oxygen.Enriched mercury isotopes were added in quantities similar to in situ concentrations of ambient Hg and MeHg (average of 47% of in situ total Hg for methylation, average of 87% of in situ MeHg for demethylation).The isotopes were injected through small silicone-covered holes spaced at 1 cm intervals along the core tubes using a 100 ml gastight syringe.Cores were then incubated at in situ temperatures for approximately 5 hours, extruded to 0-2 cm, 2-4 cm, and 4-8 cm depth intervals and ash frozen with dry ice to nish the assays.The frozen samples were freezedried and homogenized prior to analysis.The generation of enriched Me 200 Hg + and loss of enriched Me 201 Hg + were measured via ICP-MS detection.For total-Hg (Hg T ) analysis (including detection of enriched isotopes), samples were microwave digested in concentrated nitric acid, diluted with deionized water, and $0.5% by volume of BrCl was added to oxidize all Hg in the sample to Hg(II).Total mercury was quantied using a Tekran 2600 automated Hg analysis system (cold-vapour generation using SnCl 2 and gold-trap amalgamation) hyphenated to an Agilent 7700x ICP-MS.For MeHg View Article Online analysis, samples were distilled according to the methods of Horvat et al., 34 but with the addition of a different enriched MeHg spike (Me 199 Hg) for MeHg determination by isotope-dilution techniques. 35All analyses used calculations from Hintelmann and Ogrinc 36 to account for the <100% enrichment of isotopes in calculating "excess" 200 Hg and 201 Hg concentrations, as well as in calculating ambient Hg T and MeHg levels from the same sample.Methylation potential rate constants (K meth ) are reported in units of per day, while demethylation potential rate constants (K demeth ) are reported as per hour.Rather than the K meth /K demeth ratio used in some studies to examine relative methylation, 19 we multiplied K meth by the solid-phase peat inorganic Hg concentration (total mercury minus methylmercury) and K demeth by the solidphase peat MeHg concentration (to derive potential rates in units of ng g À1 d À1 ).We used the ratio of these potential rates as a measure of the "potential relative methylation rate".This measure should only be considered semi-quantitative as the bioavailable fractions of inorganic Hg and MeHg are not known or readily quantiable. 37However, the product of K meth and the solid phase inorganic mercury pool (with which it is reacting) is a more intuitive measurement than the raw K meth potential (units of d À1 ) because the product accounts for the pool of reactant.Ratios of these potential rates (each having units of ng g À1 d À1 ) above one suggest the potential for net methylation and ratios of these potential rates below one suggest the potential for net demethylation.
Porewaters were analyzed for MeHg by isotope-dilution ICP-MS following distillation, as described above for sediment samples. 34,35rewaters were analyzed for Hg T according to USEPA method 1631, using a Tekran 2600 automated mercury analyzer. 38The ratio of mean solid phase to mean porewater concentrations of MeHg and inorganic Hg at each sampling plot was used to dene an effective porewater-solid phase distribution coefficients (K D ).
Other analyses.Acid-volaltile sulde (AVS) was analyzed with the automated methylene blue method 39 and accompanying water content measurements allowed reporting on a dry mass basis.Dissolved porewater sulde was also measured using the automated methylene blue method.Filtered porewater samples were analyzed for sulfate and other major anions via ion chromatography on a Dionex ICS 1100 system.

Statistical methods
All variables were tested for normality by examination of histograms and Q-Q plots and log-transformed as necessary.Differences across seasons and between sulfate-impacted and sulfate-unimpacted sites were investigated using a mixed model taking into account wetland, sampling site, and season.Correlations were explored using Pearson's correlation.Statistical signicance was tested at a ¼ 0.05.All statistical analyses were conducted using the RStudio package. 40

MeHg distribution and production in peat
Depth variation.To contextualize the patterns of near surface versus deeper rates and concentrations, we examined depth proles across our study sites for mercury and other mercury-related variables (Fig. 2).As is the case in many freshwater wetlands (e.g., Gilmour et al. 20 ), maxima and/or signicant seasonal variation for many variables occurred within the near-surface peat, in this case within the top 4 cm.Across both sampling plots at the sulfate-impacted wetland (near open water and near upland), depth proles of MeHg concentration and % MeHg (the proportion of total Hg present as MeHg) had near-surface maxima during all seasons.However, across both sampling plots at the sulfate-unimpacted sites, MeHg concentration and % MeHg proles did not consistently exhibit near-surface maxima and were less variable with depth.Depth-proles for K meth generally exhibited surface maxima and were, on average, similar across impacted and unimpacted sites.No consistent pattern was observed for depth proles for K demeth at either site, though the minimum was typically in the surface (0-2 cm) peat.Because the important variability in processes was limited to the upper 4 cm, we focused our statistical comparisons and investigation of chemical relationships on samples pooled from 0-2 cm and 2-4 cm.
Seasonal variation.Although no signicant seasonal differences in peat MeHg or % MeHg were observed, K meth decreased signicantly (p < 0.001) between summer and fall at both sites (Fig. 3c).While K demeth was similar or decreased between summer and fall at the sulfate-unimpacted sites, K demeth increased signicantly (p < 0.001) at the impacted sites between summer and fall, which was during the intentional ooding of the wetland with sulfate-rich water and at a time when sulfate in porewater increased signicantly (Fig. 4b).These patterns generally manifested in potential relative methylation rates well above 1.0 during summer (suggesting the potential for a predominance of methylation) and less than 1.0 in fall (suggesting the potential for a predominance of demethylation).The signicant decreases in K meth and increases in K demeth observed in October at the sulfate-impacted sites resulted in the lowest potential relative methylation rates of the study.Across sites, the highest concentrations of porewater MeHg were found in July (p < 0.01; Fig. 4a).Porewater-solid phase distribution coefficients for MeHg and Hg T (Fig. 4d and e) varied signicantly both across sites and by sampling date.These signicant differences resulted from the relatively large K D values observed for MeHg in May and October in the sulfate-impacted wetland, with similar but weaker patterns observed for Hg T .
For K meth , no signicant differences were observed between sulfate-impacted and sulfate-unimpacted sites during the same season.For K demeth and potential relative methylation rates, the only signicant same-season difference between sites was observed in October when values were greater in the sulfateimpacted wetland than the unimpacted wetland (Fig. 3d and e).Similar to MeHg in peat, porewater MeHg concentrations did not differ signicantly between sulfate-impacted and unimpacted sites (p ¼ 0.12).
The mine pumping events that inundated the impacted wetland with high-sulfate water in the fall prior to our study and again from mid-September 2012 to well aer our October sampling, did not have a signicant impact on MeHg concentrations in the peat solid phase.No signicant increases in peat MeHg concentration, peat % MeHg, or porewater MeHg concentrations were observed in response to this sulfate loading, in contrast to experimental sulfate additions to sulfatelimited wetlands which produced rapid and marked increases in porewater MeHg (e.g.ref. 16, 17, 42 and 43).Instead, during July when sulfate concentrations in the impacted wetland were at their lowest (0.9-8.7 mg l À1 ; Fig. 4b), high porewater MeHg concentrations, elevated K meth , and lower pore-water-peat distribution coefficients (relative to October) suggest more rapid net methylation and relative partitioning into pore water.Although we did not observe signicantly greater sulfate reduction productssuch as porewater sulde and AVSduring July in the sulfate-impacted wetland, the large decline in porewater sulfate between May and July strongly suggests a period of active sulfate consumption by sulfatereducing bacteria.The activity of sulfate-reducing bacteria, which is oen coupled with high rates of Hg methylation, 44,45 was likely enhanced by the warmer temperatures observed during July and led to the elevated porewater MeHg, K meth and potential relative methylation rates observed at this time.The summertime maximum in MeHg production and porewater accumulation suggests that peat temperature, as it affects microbial activity, may be important to seasonal patterns.
The lack of a strong relationship between K meth or K demeth with % MeHg in the peat (p ¼ 0.16 and 0.12, respectively) most likely represents a decoupling of the timing between instantaneous Hg methylation/demethylation processes and solid phase MeHg accumulation (Fig. 5).We observed a signicant (p < 0.001), but weakly (r 2 ¼ 0.38) inverse relationship between solid phase % MeHg and potential relative methylation rates, indicating that the instantaneous methylation and demethylation potential rate constants that we measured do not reect the long-term accumulation of MeHg in these systems.This is in sharp contrast to other studies in freshwater systems, wherein % MeHg correlated positively and relatively strongly with K meth or the K meth /K demeth ratio. 19,20Our study systems, particularly the sulfate-impacted wetland, are likely not in chemical equilibriuma result of the transient hydrologic and geochemical setting, which leads to short-term swings in redox conditions in the surface peat that are difficult to discern in the seasonal sampling.The most dramatic example of this decoupling was observed during October in the sulfate-impacted wetland: solid phase MeHg concentrations were some of the highest we observed and sulfate concentrations were elevated, but potential relative methylation rates were exceedingly low.We hypothesize that net MeHg production and accumulation may have continued beyond our July sampling as temperatures increased until the wetland was inundated with high-sulfate, well-oxygenated mine-waters in mid-September.The large inux of high-sulfate, oxygenated water may have led to a redox block, such that the wetland could not attain sufficiently reducing conditions for Hg methylation, and instead Hg demethylation was favored (Fig. 3d and e) and reected in our instantaneous potential relative methylation measurements.Indeed, laboratory studies have shown that Hg methylation is coupled with sulfate reduction, but that more strongly reducing, fermentative conditions are especially conducive to Hg methylation. 46Additionally more recent research suggests that bacteria that thrive in sulfate reducing and fermentative conditions are common methylators. 10The high potential relative methylation rates observed in July during a time of low available sulfate (Fig. 3e and 4b) support this redox block hypothesis.Porewater sulde was still present in the sulfateimpacted peat in October (lower than July by a factor of 2), but porewater sulfate had increased by three orders of magnitude.Directly linking instantaneous processes with MeHg concentration over time in this sulfate-impacted system is not feasible in the eld, but laboratory experiments should be devised to more carefully trace processes across time using chronically sulfate-impacted and -unimpacted peat.

Sulfate loading impacts on Hg bioavailability
The second objective of this study was to investigate whether the chronic loading of sulfate into the Long Lake Creek wetland had a signicant impact on inorganic Hg bioavailabilty due to sulde complexation.Previous work in other sulfate-impacted freshwater systems such as the Florida Everglades suggests a "Goldilocks zone" wherein inputs of sulfate stimulate Hg methylation through the activity of sulfate-reducing bacteria, but the accumulation of sulde with further sulfate inputs and advanced sulfate reduction inhibits Hg methylation.Although several hypotheses describing the formation of microbiallyunavailable Hg-S complexes have been proposed, including mononuclear aqueous complexes, sulydryl-Hg complexes in DOC, and nanoparticulate HgS, [47][48][49] experimental observations suggest that the peak of the "Goldilocks zone" above which Hg methylation begins to be inhibited occurs at sulde concentrations between approximately 300 and 3000 ug l À1 in porewater. 47,48Results from our study, wherein we observed a strong log-linear relationship between % MeHg and porewater sulde at sulde concentrations below $600-700 mg l À1 (p < 0.01; r 2 ¼ 0.87), follow a similar pattern (Fig. 6b).This relationship excludes an outlier of low % MeHg at the near open water site in the sulfate-impacted wetland during July.At impacted and un-impacted sites collectively, we observed an equally strong relationship for AVS concentrations up to $800 mg g À1 (Fig. 6a; p < 0.01; r 2 ¼ 0.81).At the impacted site, the near open water plot had consistently higher AVS and consistently lower solid phase % MeHg, a pattern that may reect both the proximity to the sulfate-rich stream and more consistent strongly-reducing and inhibiting conditions nearer the open water.Although our observations of solid phase % MeHg are consistent with the hypothesis that MeHg production in the sulfate-impacted wetland was suppressed by low bioavailability of inorganic Hg (resulting from high sulde concentrations), Hg process measurements (K meth , K demeth , or potential relative methylation rate) did not reect this.For example, lower methylation rate constants but higher solid phase MeHg were consistently observed farther from shore at the LLC site (compared to near open water, Fig. 3a and c).The consistent relationships observed between sulde chemistry and solidphase MeHg concentration, but inconsistent relationships observed between sulde chemistry and methylation/demethylation rates, is further evidence that instantaneous Hg methylation rates are decoupled from MeHg accumulation in an environment with dynamic changes to redox conditions such as peat.Unraveling the reason for this decoupling is challenging with the limited temporal resolution of solid-phase measurements.The sampling plot near the open water in the sulfateimpacted wetland had consistently higher AVS than the sampling plot farther from shore (near upland, Fig. 3f); however, dissolved sulde (Fig. 4c) did not show a similar pattern.Potential relative methylation rate and porewater MeHg, similar to porewater sulde, likely reect conditions present in the preceding hours or days, whereas solid phase % MeHg and AVS are likely more representative of the cumulative effect of predominant microbial and redox conditions over a longer period.

Conclusions
Methylmercury production in freshwater wetlands subject to chronically elevated sulfate loading is different from observations in wetlands where sulfate is limiting.Methylmercury concentrations and methylation potential rate constants did not differ signicantly between a freshwater wetland chronically impacted by mining-inuenced water containing high (100 s of mg l À1 ) sulfate concentrations and a similar wetland with background ($5 mg l À1 ) sulfate concentrations.This was despite large apparent differences in sulfate reduction between the two wetlands, manifest in orders of magnitude differences in pore water sulfate and sulde and solid-phase AVS concentrations.Importantly, we found that short-term measurements of methylation and demethylation potentials were very weakly related to longer-term measurements of MeHg accumulation, such as %-MeHg in peat, indicating that methylation and demethylation processes are highly transient through time.While this research may be interpreted that sulfate inputs to wetland systems in this region are uncoupled from Hg methylation, we caution that our work has focused on a chronically sulfate-impacted wetland and that further research is needed to better understand the evolution of wetland Hg cycling responses during the transition from low sulfate to chronically-sulfate-impacted conditions.

Fig. 1
Fig. 1 Study site locations downstream from mining features on the Mesabi iron range in northeastern Minnesota, USA.(a) Sulfate-unimpacted site, wetland WTR.(b) Sulfate-impacted site, wetland LLC.(c) Sampling plots in wetlands, both near open water (1) and near upland (2), located at the + indicated on each wetland map.

Fig. 2
Fig. 2 Peat depth profiles of solid phase mercury and sulfide-related variables across sulfate-impacted and sulfate-unimpacted sites.Values represent the mean of three replicates; error bars have been omitted for clarity.

Fig. 4
Fig. 4 Temporal and inter-site variability in (a) porewater MeHg concentration, (b) porewater sulfate concentration, (c) porewater sulfide concentration, (d) porewater-solid phase distribution coefficient for MeHg (K D MeHg), and (e) porewater-solid phase distribution coefficient for Hg T (K D Hg T ).Values represent the mean and standard deviation of 0-2 and 2-4 cm samples.

Fig. 6 %
Fig. 6 % MeHg in peat in relation to (a) porewater sulfide and (b) acid-volatile sulfide (AVS) in peat.Solid symbols represent near open water locations and open symbols represent near upland locations in each wetland.Error bars represent the standard deviation of three replicate measurements.
The site farther from the open water may experience more frequent and prolonged dry periods which could re-mobilize sulfate from reduced solid sulfur phases providing additional fuel for driving MeHg production, as has been observed in relation to water table uctuation (Coleman Wasik et al. 2015).
50Additionally, more frequent inputs of freshwater runoff from the upland could result in lower sulfate loads to peat porewaters and less buildup of sulde in peat porewaters and solid phase.