Journal of Geophysical Research: Biogeosciences

Factors controlling temporal and spatial distribution of total mercury and methylmercury in hyporheic sediments of the Allequash Creek wetland, northern Wisconsin



[1] Hyporheic pore water samples were collected from two sites within the Allequash Creek wetland, in Vilas County, northern Wisconsin, from August 2003 to October 2004. Samples were collected simultaneously at the surface and at 2, 5, 7, 10, and 15 cm below the sediment-water interface. Concentration ranges were 3.7 to 58 pM for inorganic mercury, <0.5 to 16 pM for methylmercury, 3.02 to 152 μM and 0.38 to 96.6 μM for oxidized and reduced iron, respectively, 1.28 to 48.2 μM and< 0.05 to 9.76 μM for oxidized and reduced sulfur, respectively, and 109 to 689 μM for dissolved organic carbon. These concentrations are typical of anoxic environments such as wetlands and lake sediments. These data were used to gain a better understanding of the processes controlling spatial and temporal variability of inorganic mercury and methylmercury. Findings show that conditions conducive to mercury methylation exist in the hyporheic zone, especially in late summer, when accumulation of reduced iron and sulfide are indicative of microbial iron and sulfate reduction. Methylmercury concentrations also peak in late summer, with the highest concentrations appearing 2 to 10 cm below the sediment-water interface. While there is a general covariance of total mercury and methylmercury over the depth profile, poor correlation was observed over time, highlighting the dynamic nature of hyporheic zone conditions and suggesting changes in mercury speciation and partitioning.

1. Introduction

[2] Methylmercury (MeHg) is a highly toxic, bioaccumulative pollutant with widespread distribution throughout ecosystems. Because it is much more toxic than its inorganic form, processes controlling the production of methylmercury are of interest to ecosystem and public health managers. Wetlands are known to produce and export methylmercury to watersheds [Hurley et al., 1995; Krabbenhoft et al., 1995; St. Louis et al., 1994], primarily due to methylation of inorganic Hg(II) by sulfate- and iron-reducing bacteria in anoxic zones [Compeau and Bartha, 1985; Goulet et al., 2007; Kerin et al., 2006]. While the chemical conditions favorable to these bacteria are known, factors controlling the bioavailability of mercury to the microbial food web are not well understood.

[3] Mercury methylation occurs inside sulfate- and iron-reducing bacteria and is therefore controlled, in part, by the rate of mercury uptake by the organisms [Benoit et al., 1999]. Sulfide binding is expected to dominate mercury speciation in anoxic environments because of the high formation constants for Hg-S complexes. The net charge and the solubility of these species in turn affects their bioavailability, with neutrally charged, dissolved species being the most available [Barkay et al., 1997; Morel et al., 1998]. At sufficiently low sulfide concentrations, however, or in the presence of natural organic matter (NOM) that contains sulfide and thiol functional groups, Hg-NOM complexes can become dominant [Chadwick, 2006; Hammerschmidt et al., 2004; Hsu and Sedlak, 2003]. The exact impact of Hg-NOM complexation on bioavailability is unclear. Natural organic matter complexes of Hg(II), in which Hg(II) is bound to sulfur-containing functional groups, are expected to be unavailable to bacteria, yet in many systems, MeHg concentrations correlate positively with NOM concentrations, suggesting a more complex relationship between organic ligand binding and bioavailability [Babiarz et al., 1998; Hammerschmidt and Fitzgerald, 2004; Hammerschmidt et al., 2004]. The abundance of these Hg(II) complexes is dependent on both the redox conditions and the chemistry of natural waters.

[4] The hyporheic zone of streambeds is a subsurface, three-dimensional region in which the active exchange of oxidized surface water and reduced groundwater produces sharp chemical gradients with depth [Alley et al., 2002]. Typically, the water in this zone exchanges between the surface and subsurface many times along its flow path, influencing the chemistry of both the surface and groundwater. The exchange of oxic and anoxic water, along with high concentrations of labile organic matter, often gives rise to substantial biological activity that may support mercury methylation and transport into both surface and groundwater [Bencala, 2000; Harvey and Wagner, 2000; Winter et al., 1998].

[5] In this paper, we characterize the hyporheic zone of the Allequash Creek wetland ecosystem in terms relevant to the bioavailability of mercury to methylating bacteria and determine whether there is evidence of in situ methylmercury production. We focused on depths up to 15 cm below the sediment-water interface, fully encompassing the region of significant surface/groundwater mixing [Kerr, 2007; Kerr et al., 2008; Schindler and Krabbenhoft, 1998]. We compare a suite of biogeochemical measurements from the hyporheic pore waters taken at two sites over 14 months in order to characterize the processes controlling seasonal variations in methylmercury concentrations in the context of changing redox conditions and hyporheic zone chemistry.

2. Site Description

[6] The Allequash Creek watershed is located in the Flambeau River Basin, in Wisconsin's Northern Highland Lake District (Figure 1). The creek is spring-fed, originating in a small beaver pond at the eastern end of the watershed. It flows 4 km through a riparian wetland before discharging into Allequash Lake and subsequently into Trout Lake [Kerr, 2007]. The wetland has peat soil exceeding 6 m in depth [Elder et al., 2000]. Its plant community is dominated by sphagnum moss, leatherleaf, tussocks sedge, and black spruce [Krabbenhoft et al., 1995; Magnuson et al., 1984; Schindler and Krabbenhoft, 1998].

Figure 1.

Map of the study area showing the locations of the sampling sites. Middle wetland site located at 46.03177°N, 89.60762°W; upper springs site located at 46.03433°N, 89.58030°W. Wetlands, as delineated in the National Land Cover Database [Multi-Resolution Land Characteristics (MRLC) Consortium, 2001], are shown in light gray. Contour interval: 10 m.

[7] The hydrology and biogeochemistry of the Allequash Creek watershed is well-characterized, both through the U.S. Geological Survey's Water, Energy, and Biogeochemical Budgets (WEBB) program [Elder et al., 1992; Krabbenhoft et al., 1995; Walker and Bullen, 2000], and the National Science Foundation's North Temperate Lakes Long-term Ecological Research (NTL-LTER) program [Callahan, 1984; Magnuson et al., 1984].

[8] More than 90% of streamflow in Allequash Creek is derived from groundwater [Pint et al., 2003]. The surface watershed above the discharge point of Allequash Creek into Allequash Lake, is 9.9 km2, while the groundwater basin to the same point is 15.0 km2 [Krabbenhoft et al., 1995]. The soils in the region are moderately hydraulically conductive (average K ≈ 10−5 m/s, horizontal K = 10 m/d [Hunt et al., 2006]), and the average groundwater recharge rate is 27 cm per annum [Hunt et al., 1998]. Average precipitation in the area is 79 cm per annum [Peters et al., 2006].

[9] Samples were collected at two contrasting sites in the Allequash Creek watershed, in order to characterize wetland processes in areas with different groundwater flow regimes and soil chemistry. The upper springs site is located in the spring-fed headwaters of the creek, an area characterized by sandy, predominantly mineral sediments, relatively high concentrations of oxidized sulfur compounds, and low concentrations of hydrogen sulfide. Groundwater discharging to this area has typically traveled along short flow paths and is therefore relatively young (∼25–50 years). The middle wetland site has comparatively longer groundwater flow paths (and therefore older groundwater), more organic-rich sediments, a higher percentage of reduced sulfur, and higher iron concentrations, indicating more reducing conditions than at the upper springs site [Pint et al., 2003].

3. Field Methods

[10] All sampling and analysis was carried out following trace metal clean techniques [Gill and Fitzgerald, 1985; Hurley et al., 1996; Shafer et al., 1999]. Sample containers were rigorously cleaned, and the “Clean Hands, Dirty Hands” protocol [Fitzgerald, 1999] was followed to ensure sample integrity.

[11] Hyporheic pore water samples were collected using Differential In situ Pore water Samplers (DIPS), which were modeled after the MINIPOINT sampler, developed by Duff et al. [1998]. These consisted of 5 hollow Kynar® rods threaded through holes drilled in a 1 inch thick acrylic plate, in a circular arrangement (Figure 2). The rods each had a 1 cm screen length and sampled at depths of 2, 5, 7, 10, and 15 cm, respectively. The sampler was held in position with the aid of a tripod. Teflon tubing was connected to each rod, which, in turn, connected to C-Flex® tubing at the pump head. Water was drawn through the samplers using a multichannel peristaltic pump capable of pulling water from all of the rods simultaneously at a rate of 1–2 mL/min. This slow rate of withdrawal minimized the cone of depression around each sampler inlet and allowed for high spatial resolution throughout the depth profile. The pump head tubing discharged into a 0.45 μM pore size polysulfone membrane filter, then through another length of Teflon tubing and into a 500 mL Teflon bottle. Two of these samplers were used at each site, and the water from the two samplers was collected into a single, composite container for each depth, in order to minimize differences due to spatial and temporal variation and to ensure a more representative sample. Aliquots for mercury, total iron, and sulfur analysis were subsampled from the composite, to minimize sample variation over the collection period. Samples for redox-sensitive analytes (Fe(II) and sulfide) were collected directly from the pump discharge tubing to minimize exposure to air [Duff et al., 1998; Kerr, 2007; Meyer, 2005].

Figure 2.

Schematic of the Differential In situ Pore water Sampler [Meyer, 2005].

4. Analytical Methods

4.1. Hg and MeHg

[12] Samples for mercury and methylmercury analysis were preserved in the field with 6N TraceMetal Grade™ HCl to a final concentration of 1%, and were stored in a 4.5°C cold room until analysis. Analytical methods are described in detail elsewhere [Babiarz et al., 1998; Bloom and Crecelius, 1983; Bloom, 1989; Horvat et al., 1993; Liang et al., 1994]. Briefly, samples for total mercury (HgT) analysis were digested overnight with BrCl oxidant, reduced with SnCl2, and analyzed by dual gold amalgamation Cold Vapor Atomic Fluorescence Spectrometry (CVAFS), using a Tekran 2500 analyzer. Methylmercury samples were distilled to remove matrix interferences, ethylated using sodium tetraethylborate, and purged onto Carbotrap® columns. The mercury was then thermally desorbed from the columns into an analytical train consisting of an isothermal chromatographic column, a pyrolytic column, and a Tekran 2500 CVAFS analyzer. All species were quantified as Hg(0). The limit of detection was typically below 10 pg.

4.2. Iron

[13] Ferrous iron was measured spectrophotometrically in samples that were fixed in the field by pouring into tubes containing ferrozine reagent [Stookey, 1970]. Total iron was measured using a Thermo-Finnigan Element2 magnetic sector inductively coupled plasma-mass spectrometer (ICP-MS). Ferric iron concentrations were calculated as the difference between total and ferrous iron.

4.3. Sulfide and Sulfate

[14] Sulfide was measured using an ion-selective electrode in samples that were fixed in the field by pouring into tubes containing Sulfide Anti-Oxidant Buffer. The limit of detection was 0.05 μM. Sulfate was measured by ion chromatography and ICP-MS in samples acidified to 0.25 N with nitric acid [Kerr, 2007].

4.4. Dissolved Organic Carbon (DOC)

[15] DOC was measured using a Shimadzu TOC-V CSH/CSN Total Organic Carbon analyzer. Analytical precision was typically 1–4% [Kerr, 2007].

5. Results

[16] Iron and sulfur profiles show that reducing conditions were present below 2 cm in the hyporheic zone, as indicated by increased concentrations of reduced iron and sulfide (Figure 3). Fe(II) concentrations peak lower in the profile at the upper springs (15 cm) than at the middle wetland (2 cm), illustrating the contrast in reducing conditions between the two sites [Kerr, 2007]. The sulfate and sulfide profiles follow a similar pattern, with sulfide as a percentage of sulfate peaking at 15 cm at the upper springs and reaching a value very close to the peak at 2 cm at the middle wetland.

Figure 3.

Representative summer sulfur and iron profiles from the 2003–2004 sampling period. Upper figures are from the upper springs site, lower figures are the middle wetland site. Modified from Kerr [2007].

[17] Total mercury concentrations in the pore water were generally higher than those in surface water (Figure 4). At the upper springs site, they remain relatively high from 2 cm to 7 cm (∼25 pM), then drop off rapidly to a level close to the surface water value. The middle wetland total mercury profile displays a similar trend between 2 and 7 cm with a concentration of roughly 29 pM, but the concentration jumps even higher at 10 cm (∼36 pM), before tapering to an intermediate concentration (∼22 pM) at 15 cm. Pore water total mercury concentrations were higher at the middle wetland than at the upper springs in every subsurface measurement. Methylmercury concentrations at both sites peak at 7 cm, with a larger peak at the middle wetland than at the upper springs. Concentrations at both sites drop to their lowest levels at 15 cm, the deepest point sampled.

Figure 4.

Representative late summer depth profiles of total mercury and methylmercury. Upper springs profile collected 27 August 2003, middle wetland profile collected 26 August 2003.

[18] Total mercury concentrations at the middle wetland vary seasonally, with peak concentrations in the late summer and early fall, and also vary considerably with depth during the sampling period (Figure 5). Methylmercury concentrations and the percentage of total mercury represented by MeHg were consistently higher in the early fall than in summer (Figures 6 and 7) . MeHg concentrations varied less with depth than total mercury concentrations. Dissolved organic carbon concentrations follow a similar pattern to methylmercury in every sample except the surface water (Figure 8).

Figure 5.

Total mercury concentrations by depth in the hyporheic zone at the middle wetland site during the 2003–2004 sampling period.

Figure 6.

Methylmercury concentrations by depth in the hyporheic zone at the middle wetland site during the 2003–2004 sampling period.

Figure 7.

Methylmercury as a percentage of total mercury by depth in the hyporheic zone at the middle wetland site during the 2003–2004 sampling period.

Figure 8.

Dissolved organic carbon concentrations by depth in the hyporheic zone at the middle wetland site during the 2003–2004 sampling period.

6. Discussion

[19] The hyporheic zone provides geochemical conditions conducive to methylation of inorganic Hg(II) in the summer and early fall months. The presence of reduced sulfur and iron species indicates that sulfate-and iron-reducing bacteria are active in the sediments (Figure 3). Organic matter, likely derived from both the hyporheic sediments and stream water, serves as a carbon source for microbial respiration, resulting in anoxic conditions. The Fe(III) required to sustain iron-reducing bacteria is most likely derived through the reductive dissolution of iron oxyhydroxides in the hyporheic sediments, while the sulfate on which sulfate-reducing bacteria depend is derived from surface water [Kerr, 2007]. The different sources of these two electron acceptors explain the differences in the observed profiles of Fe(II) and sulfide. The sediments are also enriched in inorganic Hg(II) [Meyer, 2005], apparently through retention of Hg(II) transported into the hyporheic zone by groundwater and/or surface waters. The sediments then provide a source of inorganic Hg(II) for methylation (Figure 4), presumably by sulfate- and/or Fe(III)-reducing bacteria in the pore water [Compeau and Bartha, 1985; Goulet et al., 2007; Kerin et al., 2006]. The appearance of MeHg in the middle and upper hyporheic zone, especially during the fall season, at concentrations higher than either the groundwater or the surface water, indicates that in situ methylation, rather than transport, is the primary source. The concentrations and proportions of MeHg are typical of anoxic environments such as wetlands and lake and marine sediments [Goulet et al., 2007; Hammerschmidt and Fitzgerald, 2004; Hines and Brezonik, 2007; Hurley et al., 1995].

[20] The hyporheic zone is a dynamic system in terms of both biogeochemical and hydrologic processes. Thus, several factors could account for the observed temporal and spatial variability in concentrations of both HgT and MeHg. Partitioning of mercury between hyporheic sediments and pore waters likely exerts a strong influence on the concentrations observed in the pore water. Variations in the rates of flushing of the hyporheic zone by groundwater and surface waters may result in non-equilibrium conditions and variability in concentrations. Although concentrations of Fe(II) and sulfide are indicative of microbial reduction of Fe(III) and sulfate, processes associated with methylation, the activities of the associated microbial communities are known to vary, depending on hyporheic zone conditions [Gilmour et al., 1992; Hammerschmidt and Fitzgerald, 2004; Kerin et al., 2006]. Changes in microbial methylation rates are likely major drivers of the observed seasonal trends in MeHg levels (Figures 6 and 7), but could also account for some of the variation with depth during the fall season, when the highest levels are observed. The measured concentrations of MeHg are the result of net methylation, and therefore, rates of demethylation likely influence the concentrations of MeHg in pore waters of the hyporheic zone [Goulet et al., 2007; Hammerschmidt and Fitzgerald, 2004]. While demethylation rates have been shown to be insignificant in some systems [Goulet et al., 2007], in others, they can play a major role in net methylation [Hintelmann et al., 2000]. In follow-up studies, direct measurements of demethylation rates are being made in this system, in order to determine the influence of this process on MeHg cycling and profile structure.

[21] If mercury methylation were controlled solely by the presence of reducing conditions, the availability of an electron acceptor, a carbon source, and inorganic mercury, we would expect methylmercury concentrations to correlate with concentrations of reduced iron and sulfur. The lack of a clear trend in the methylmercury depth profile (Figure 4) to match the highly variable Fe(II) and sulfide gradients (Figure 3), however, suggests that at the time this profile was collected, another factor, such as the speciation of inorganic mercury, was also playing a role, by controlling the bioavailability of inorganic Hg(II) to Fe(III)- and/or sulfate-reducing bacteria. In particular, binding of inorganic mercury to sulfide and/or DOC, both of which are present at relatively high levels (Figures 3 and 8), may be influencing methylation rates.

[22] Our speciation modeling predicted that HgS0 should be the dominant species of Hg(II) in sulfidic (hyporheic) waters, while Hg-NOM complexes should be dominant in the absence of sulfide [Armstrong et al., 2006] (in contrast, for MeHg, both MeHgS0 and MeHg-NOM complexes were predicted to be important in the presence of sulfide). Experimental and theoretical evidence suggests that the neutral species HgS0 can readily traverse biological membranes, thus the bioavailability of Hg(II) should be high in the sulfidic hyporheic zone [Benoit et al., 1999]. However, experimental measurements by our group indicated that Hg-NOM complexes were predominant in both sulfidic and non-sulfidic waters [Chadwick, 2006]. While confirmation of these direct speciation measurements is needed, apparently, there are either kinetic constraints to reaching equilibrium in dynamic, anoxic systems, or the binding constants used in models are inaccurate. Thus, the role of speciation in controlling methylation rates and accumulation of MeHg is uncertain.

[23] Because sulfide and NOM concentrations in this system vary within a relatively narrow range, large shifts in Hg speciation are not expected, therefore, one likely explanation for the seasonal variation in methylmercury and DOC concentrations (Figures 68) is higher activity of the microbial community when water temperatures are warmer, allowing for a higher rate of organic matter decomposition and mercury methylation. Methylmercury concentrations at most depths (Figure 6) peak shortly after the peaks in DOC concentrations (Figure 8), which may reflect the use of some of the DOC as a carbon source for the sulfate- and iron-reducing bacteria. This is consistent with the previous finding from this site that aliphatic DOC produced in the hyporheic zone serves as an energy source for organisms close to the sediment-water interface [Schindler and Krabbenhoft, 1998].

[24] The seasonal variation in total mercury concentrations is most likely due to two factors. The first is changes in the rate of flushing of the hyporheic zone by groundwater and/or surface waters [Meyer, 2005; Kerr, 2007]. Because these waters have low mercury concentrations, we expect the net effect of increased flushing to be a reduction in pore water dissolved Hg(II) concentrations. The second factor is changes in solubility, due to altered sediment-water partitioning driven by redox conditions. Concentrations of important ligands for Hg(II) and MeHg, especially sulfide and DOC, drive the partitioning of mercury species, and thus their concentrations in pore water.

[25] The difference in subsurface total mercury concentrations between the two sites can be explained, at least in part, by differences in hydraulic conductivity. The sandy sediments of the upper springs site have a higher hydraulic conductivity and, presumably, because of their lower organic content, a lower binding affinity for dissolved inorganic mercury than the organic-rich sediments of the middle wetland site. This difference in total mercury concentrations may explain why the upper springs methylmercury depth profile peaks at a lower concentration than the middle wetland profile. The upper springs site has higher methylmercury concentrations at some depths than the middle wetland site, however, illustrating the high spatial variability that is common in this area. Several differences between the two sites are likely to affect the ability of the microbial community to methylate mercury. At the upper springs site, groundwater influencing the upper regions of the core has traveled along a shorter flow path than at the middle wetland site, and therefore has higher dissolved oxygen concentrations [Kerr, 2007; Pint et al., 2003]. Also, sulfate-rich surface water is expected to penetrate deeper into the hyporheic zone at the upper springs site, due to its higher hydraulic conductivity. While the sulfate is expected to enhance anaerobic microbial respiration, the dissolved oxygen should have an inhibitory effect. In contrast, the hydrology of the middle wetland hyporheic zone is dominated by the upwelling (discharging) of anoxic, older groundwater. This factor, coupled with reduced entrainment of both oxygen and sulfate from surface waters, results in a hyporheic zone with oxygen and sulfate concentrations lower at each depth than at the upper springs site.

7. Conclusions

[26] Bacterial activity, inorganic mercury concentrations, and mercury speciation are believed to be primary factors controlling the methylation of inorganic Hg(II). Speciation of Hg(II) is expected to be important by controlling the bioavailability of inorganic mercury to sulfate- and iron-reducing bacteria. This investigation of hyporheic zone sites in the Allequash Creek wetland was conducted to examine the levels of total- and methylmercury in hyporheic zone pore waters and their temporal and spatial variability. The hyporheic zone was chosen because it provides conditions favorable to sulfate- and iron-reducing bacteria and variable levels of Hg(II)-binding ligands.

[27] We found that while favorable conditions exist for the methylation of mercury (reducing conditions, labile carbon supply, electron acceptors, inorganic mercury), these conditions vary both spatially and temporally. Our findings show that methylmercury concentrations also vary spatially and temporally, seemingly in response to seasonal variations in conditions that influence the activity of bacteria responsible for methylation. Levels of Hg(II)-binding ligands, especially sulfide and DOC, may also be important. Subsequent work will examine methylation rates and the influences of speciation.

[28] Our findings are important to biogeochemists and ecosystem managers alike because they highlight the importance of linking the levels and distribution of total and methylmercury to biogeochemical conditions. Through understanding these conditions and processes, improved predictions can be made of the potential for methylmercury production among diverse ecosystems.


[29] Joel Overdier and Jackson Helmer provided critical field and laboratory support for this research. This work was funded in part by generous grants from the Wisconsin Groundwater Coordinating Council and the University of Wisconsin Water Resources Institute (project IDs WR04R001 and WR07R008). Funding was also provided by the United States Environmental Protection Agency (EPA) under the Science to Achieve Results (STAR) Graduate Fellowship Program (project ID FP-91687801). EPA has not officially endorsed this publication, and the views expressed herein may not reflect the views of the EPA.