Potential demethylation rate determinations in relation to concentrations of MeHg, Hg and pore water speciation of MeHg in contaminated sediments

Marine Chemistry 112 (2008) 93–101

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Marine Chemistry j o u r n a l h o m e p a g e : w w w. e l s ev i e r. c o m / l o c a t e / m a r c h e m

Potential demethylation rate determinations in relation to concentrations of MeHg, Hg and pore water speciation of MeHg in contaminated sediments Andreas Drott a,⁎, Lars Lambertsson b, Erik Björn b, Ulf Skyllberg a a b

Department of Forest Ecology and Management, Swedish University of Agricultural Sciences, S-90183 Umeå, Sweden Department of Chemistry, Umeå University, S-90187 Umeå, Sweden

a r t i c l e

i n f o

Article history: Received 30 January 2008 Received in revised form 19 June 2008 Accepted 7 July 2008 Available online 12 July 2008 Keywords: Methyl mercury Estuaries Pulp wastes

a b s t r a c t Specific, potential demethylation rate constants (kd, day− 1) were determined in fresh and brackish water sediments from seven different sites in Sweden originally contaminated with either Hg0(l) or phenyl-Hg. Variations in kd among and within sites were related to ambient concentrations of Hg (1–1143 nmol g− 1) and MeHg (4.4–575pmol g− 1), and to pore water speciation of MeHg. Chemical speciation modeling revealed that MeHgSH(aq), MeHgS−(aq) and MeHg–thiol complexes [MeHgSR(aq)] associated to dissolved organic matter were the dominant MeHg species in the sediment pore water at all sites. Potential rates of MeHg demethylation were determined as the decomposition of isotopically enriched Me204HgCl during 48 h of incubation in darkness under N2(g) at 23 °C. There was a significant (p b 0.001) positive relationship between ambient MeHg concentrations in sediments and kd across all sites, but no significant relationship between ambient Hg and kd. At the three sites with the highest ambient Hg concentrations in sediments (average ± SD, 185 ± 249 nmol g− 1), kd was not significantly correlated with pore water MeHg speciation. At sites with lower concentrations of ambient Hg in sediments (average ± SD,11 ± 8.4 nmol g− 1), there was a significant (p = 0.02) positive relationship between calculated concentrations of MeHgSH(aq), MeHgS−(aq), or the sum of these two species, and kd. If it is assumed that an oxidative demethylation process dominated at sites with lower concentrations of ambient Hg in sediments, the results suggest that it may be dependent on a passive uptake of inorganic MeHgSH molecules. It was shown that additions of different amounts of MeHg and Hg tracers, in relation to the ambient concentrations of MeHg and Hg, could result in dramatically different kd values within and between sites. At one brackish water site, both absolute demethylation rates and kds were significantly, inversely related to ambient concentrations of MeHg (and Hg). In contrast, at another brackish water site with generally less kds, samples with low ambient MeHg experienced toxic effects and demethylation was not detected. This implies that added (and possibly ambient) MeHg/Hg, depending on the environmental conditions, may have either stimulating or inhibitory effects on demethylation processes. © 2008 Elsevier B.V. All rights reserved.

1. Introduction Biomagnification of mono methyl mercury (MeHg) in aquatic food webs has resulted in high concentrations of this

⁎ Corresponding author. Tel.: +46 90 7868546; fax: +46 907868163. E-mail addresses: [email protected] (A. Drott), [email protected] (L. Lambertsson), [email protected] (E. Björn), [email protected] (U. Skyllberg). 0304-4203/$ – see front matter © 2008 Elsevier B.V. All rights reserved. doi:10.1016/j.marchem.2008.07.002

organomercury compound in fish. To better manage this problem, it is important to understand processes that control the concentration of MeHg in sediments, most importantly MeHg production (methylation) and MeHg degradation (demethylation). In recent research, much of the focus has been on methylation and its dependence on Hg speciation in pore water [e.g. Benoit et al., 2001; Benoit et al., 1999a,b; Drott et al., 2007b]. To our knowledge, no report has yet been published in which the influence of pore water MeHg speciation on demethylation has been examined.

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Demethylation of MeHg can proceed through biotic or abiotic mechanisms. Suggested abiotic mechanisms include photodegradation (Sellers et al., 1996) and degradation via formation of an unstable complex between MeHg and S2−, eventually forming volatile Me2Hg and HgS (Deacon, 1978). Experiments conducted under presumably sterile conditions have indicated that abiotic mechanisms could contribute significantly to demethylation reactions in estuarine sediments. However, the sterilization procedure used may not have been effective enough, and further investigations are therefore required (RodríguezMartín-Doimeadios et al., 2004). Principal biotic pathways for MeHg degradation are oxidative demethylation, and reductive MeHg degradation via the mercury resistance (mer) operon. During oxidative demethylation, the methyl group (CH−3) is oxidized to CO2. Oxidative demethylation is not believed to be an active detoxification mechanism, but is suggested to contribute to microbial metabolism (Marvin-DiPasquale and Oremland, 1998). Oxidative demethylation has been observed in both estuarine and freshwater sediments through the activity of aerobic as well as anaerobic organisms (Oremland et al., 1991). In contrast to the oxidative pathway, reductive mermediated MeHg degradation is an active detoxification mechanism that is induced via the mer operon. To detoxify MeHg and other organomercury compounds (broad-spectrum resistance), the enzyme organomercurial lyase, encoded for by the merB gene, cleaves the carbon–Hg bond, and the enzyme mercuric reductase, encoded for by the merA gene subsequently reduces HgII to Hg0 (Robinson and Tuovinen, 1984). The merP and merT genes are responsible for an active transport of Hg(II) into the cell (Silver and Misra, 1988). However, this transport system is not involved in transport of MeHg (Kiyono et al.,1995). Instead, it has been proposed that MeHg diffuses passively into the cell as a neutral compound (Barkay et al., 2003). The mer operon is often located on plasmids, making dissemination throughout microbial communities possible. The operon is widespread in the environment (Liebert et al., 1999), occurring in soils (Pearson et al., 1996) as well as in marine systems (Dahlberg and Hermansson, 1995). The mer operon has been demonstrated to be induced by Hg(II) or phenyl mercury acetate [e.g. Nucifora et al., 1989]. To the best of our knowledge, however, induction of the mer operon by MeHg has not been unequivocally demonstrated. Apart from mer-induced reductive demethylation, a reductive MeHg degradation mechanism in which HS−, produced by sulfate reducing bacteria, reacts with two MeHg to finally form HgS(s), MeHg and CH4, has also been proposed (Baldi et al., 1993). Because of advances in analytical methodologies, it is possible to determine the rate of demethylation using isotopically labeled MeHg. With such methods, reductive MeHg degradation via the mer operon has been found to dominate in severely Hg contaminated environments (total Hg concentration range in sediments 22–106 nmol g− 1), while oxidative demethylation has been found to dominate under less contaminated conditions (total Hg concentration range in sediments 0.01–63 nmol g− 1) (Marvin-DiPasquale et al., 2000; Schaefer et al., 2004). The threshold of Hg and/or MeHg concentrations required for reductive mer-mediated demethylation to dominate remain uncertain (Benoit et al., 2003). In this study, seven sediments with different levels of Hg contamination, salinities and redox conditions were used to

examine the influence of ambient Hg and MeHg, and MeHg pore water speciation, on laboratory determined specific, potential demethylation rate constants (kd, day− 1). In addition, the influence of added MeHg and Hg tracers on determined demethylation rates and kd was examined in two brackish water sediments. 2. Materials and methods 2.1. Site descriptions In 2004–2005, sediment samples were collected at seven contaminated sites in Sweden. Three sites had been contaminated by Hg0(l), mainly from chlor-alkali industry; Köpmanholmen (Köp), Skutskär (Sku) and Marnästjärn (Mar), and four sites had been contaminated by phenyl-Hg from pulp and paper industry; Turingen (Tur), Övre Svartsjön (Sva), Nötöfjärden (Nöt) and Karlshäll (Kar). The ambient total Hg concentration was about one order of magnitude higher at the sites that had been contaminated by Hg0(l) (Table S1, Supplementary information). Thus, the sites could be divided into two sub-sets: 1) sites contaminated by Hg0(l) having relatively high ambient Hg concentrations (Köp, Sku, Mar, average total Hg concentration, ± SD, 185 ± 249 nmol g dw− 1), and 2) sites contaminated by phenyl-Hg with lower ambient Hg concentrations (Tur, Sva, Nöt and Kar, average total Hg, ± SD, 11 ± 8.4 nmol g dw− 1). Thermodesorption (TD) measurements (Biester and Scholz, 1997), showed that Hg(II) was the dominant form of Hg at all sites, with only traces of Hg0 remaining at site Köp (data not shown). Phenyl-Hg is known to be unstable in the environment, and therefore degrade to Hg(II). With the exception of site Mar, all sites had been subjected to pulp–fibre discharge. The sites were chosen to represent gradients in salinity, annual air temperature sum, organic matter content and C/N ratio (Table S1, Supplementary information). Concentrations of chloride varied between 90 mM at the brackish water sites and 0.2 mM in freshwater lakes (Table S2, Supplementary information). Differences in air temperature and C/N ratio reflect a gradient in primary productivity (Drott et al., 2007b), with the productivity being greatest at the southern freshwater sites (Mar, Tur, Sva and Nöt), intermediate at the brackish water sites (Sku and Köp) and lowest at the northern freshwater site (Kar). The brackish water sites represented rather large estuaries, while the freshwater sites represented either small lakes (Tur and Sva) or smaller estuaries, sheltered from the open sea (Nöt and Kar). 2.2. Sampling and sample treatment Polycarbonate core samplers (diameters 70 and 80 mm) were used for sediment sample collection, and at each sampling point sediment cores (n = 5–15) were sampled within approximately 1 × 1m plots. The cores were immediately sectioned by depth (in 3, 5 or 10 cm layers), and sections from the same depth were pooled in N2-flushed plastic buckets that were filled to the top. The buckets were tightly sealed and were transported on ice to the laboratory. In the laboratory, the pooled sediment samples were homogenized. After homogenization, pore water was extracted by centrifugation followed by filtration (0.45 µm) of the supernatant. Sub-samples were taken for determination of demethylation rates, and total concentrations of Hg, C and N. Oxidation of

A. Drott et al. / Marine Chemistry 112 (2008) 93–101 Table 1 Average pore water species distribution of MeHg (±SD) for sites and sub-sets of sites Site and depth (cm)

N

% MeHgSR

% MeHgSH

% MeHgS−

Köp 0–20, 0–25, 0–40 Köp 0–10 Sku 0–25, 0–100 Sku 0–10

16 16 17 3(5)

Kar 0–20 Southern freshwaters b 0–10

10 19

17 ± 23 0.2 ± 0.2 18 ± 25 0.2 ± 0.1 (27 ± 37) a 8±7 17 ± 19

54 ± 24 88 ± 3 55 ± 22 41 ± 34 (31 ± 28) 90 ± 7 73 ± 19

29 ± 12 11 ± 3 27 ± 20 59 ± 34 (42 ± 34) 2±1 10 ± 6

Sampling depths are given for each profile in cases with several sediment profiles. a Parentheses refer to results where dredged disposed sediment is included. b Sites Mar + Tur + Sva + Nöt.

sediments and pore waters in the laboratory was minimized by working in a glove-box, under a N2(g) atmosphere. pH and dissolved H2S concentrations were measured on site, as well as in the laboratory. A more detailed description of sampling and sample treatment is given in Drott et al., 2007a.

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These assumptions are based on results from combined Hg EXAFS and S XANES studies of organic soils and of DOC from organic soils (Qian et al., 2002; Skyllberg et al., 2006). The concentration of sulfate [SO2− 4 ] was calculated as [total S]– [H2S]–[HS−]–[total organic S], where total organic S was estimated to be 0.7mass% of DOC (Qian et al., 2002). For the few samples in which [H2S] was below the detection limit (0.3 µM), the detection limit concentration was used in chemical speciation calculations. All speciation calculations were performed at a temperature of 25 °C. The ionic strength in pore waters differed among sites and chemical activities were calculated using the Davies equation. The following formation constants were used: log K1 MeHgSR = 16.5 (Karlsson and Skyllberg, 2003), log K1 MeHgSH = 14.5 (Dyrssen and Wedborg, 1991), log K1 MeHgCl = 5.25 (Schwarzenbach and Schellenberg, 1965), log K1 MeHgOH = 9.37 (Schwarzenbach and Schellenberg, 1965). For the reaction MeHgSH = MeHgS− + H+ a pKa of 7.5 (Dyrssen and Wedborg, 1991) was used. pKa values for RSH and

2.3. Chemical analyses Concentrations of MeHg in pore water were determined by species specific isotope dilution (SSID) analysis, using an enriched (97.7%) Me202HgCl-standard (Snell et al., 2004), which was equilibrated with the pore water for at least 24 h prior to analysis. The samples were ethylated using NaBEt4, and derivatised MeHg was purged and trapped on Tenax adsorbent columns (Lambertsson and Björn, 2004). Ethylated MeHg was desorbed from the Tenax adsorbent columns onto a GC–ICPMS system (Agilent 7500 ICPMS, Agilent 6890N GC) (Larsson et al., 2005). The concentration of the Me202HgClstandard used was determined by reverse isotope dilution, using a natural isotopic abundance MeHgCl-standard (Riedelde Haen). The method detection limit for MeHg in pore water was 0.02 fmol g− 1 (n = 10), and the precision was determined to be 3% relative standard deviation (RSD, n = 5) (Lambertsson and Björn, 2004). Pore water DOC and DIC were analyzed with a Shimadzu TOC-5000 analyzer. Pore water Cl and Br were analyzed using anion-exchange HPLC with conductivity detection (Dionex 4000i), and pore water total S was analyzed by ICPMS (PerkinElmer Elan 6100 DRC). Total Hg in sediments was measured by solid combustion atomic absorption spectrometry on a LECO AMA 254 mercury analyzer. The accuracy of the total Hg measurements was regularly verified by analyzing marine sediment certified reference materials MESS-2 (National Research Council of Canada) and IAEA356 (International Atomic Energy Agency), respectively, at random positions in the sample queue. Total C and N concentrations in dried, ground, homogenized sediment samples were measured on a PerkinElmer 2400 CHN elemental analyzer. 2.4. Pore water chemical speciation calculations Concentrations of RSH (organic thiol groups associated with DOC) were calculated assuming that RSH comprised 30% of reduced S and that reduced S comprised 0.5mass% of DOC.

Fig. 1. Relationship between the concentration of HS− and the % MeHgSR of total pore water MeHg for; a) brackish sediments (16–882 mg DOC L− 1, corresponding to 0.8–41 µM of RSH), b) freshwater sediments (9–38 mg DOC L− 1, corresponding to 0.4–1.8 µM RSH) and c) comparison of brackish and freshwater sediments in the low µM concentration range of HS−.

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H2S were set to 9.96 and 7.0, respectively, in agreement with Karlsson and Skyllberg (2003) and Stumm and Morgan (1996). 2.5. Potential MeHg demethylation rates and ambient sediment MeHg determinations Potential MeHg demethylation rates were determined in laboratory experiments by spiking sediments with an aqueous solution of isotope enriched Me204HgCl (98.11%). For accurate determination of the sediment Me204Hg concentration at the

start of the incubation, the samples were divided in two parts, of which one was immediately frozen at − 20 °C (t = 0 days, t0). The other part was placed in an N2(g) flushed glove-box and was incubated in darkness at 23 °C for 48 h (t = 2 days, t2), after which incubation was stopped by freezing at − 20 °C. Because the same samples were used for methylation rate determinations (Drott et al., 2007b), isotope enriched 201Hg(NO3)2 was also added to the samples. Prior to sample preparation, thawed sediment samples were spiked with an aqueous solution of Me200Hg+ to 0.5% of

Fig. 2. Relationships between added tracers of Me204Hg (a) and 201Hg (b) and the determined potential specific demethylation rate constant kd (day− 1), relationships between ambient MeHg and kd (c) and demethylation rate (d), and relationships between ambient Hg and kd (e) and demethylation rate (f). Filled circles and solid lines represent data and data fits at site Sku (0–100 cm, n = 16) and open triangles and dotted lines represent data and data fits at site Köp (0–40 cm, n = 16). Relationships having R2 values exceeding 0.54 are significant at the p b 0.001 level. R2 = 0.46, p b 0.003, R2 = 0.37, p b 0.01.

A. Drott et al. / Marine Chemistry 112 (2008) 93–101

the Hgtot concentration for SSID calibration. In addition, an aqueous solution of 199Hg2+ was added to the samples at 40% of the Hgtot concentration to correct for sample preparation MeHg artifacts. Demethylation of Me204Hg during the incubation was calculated as the difference between the t0 and t2 Me204Hg concentrations, which were solved from the measured 202/204 isotope ratio by reverse isotope dilution calculation, based on the determined ambient MeHg concentration. The concentration of all isotopic standards used was controlled by reverse isotope dilution, using natural isotopic abundance MeHgCl and HgCl2 aqueous standards (MeHgCl Pestanal grade, Riedel-de Haen and HgCl2 99.999%, SigmaAldrich). MeHg was solid–liquid extracted using a mixture of KBr/CuSO4/H2SO4/CH2Cl2, derivatised with NaBEt4 (Lambertsson et al., 2001) and analyzed by GC–ICPMS (Agilent 7500 ICPMS, Agilent 6890N GC) (Larsson et al., 2005). The method precision for total MeHg concentrations in sediments and potential MeHg demethylation rate determinations was 3% relative standard deviation (RSD), based on replicate subsample incubations and analyses (n = 9). The method detection limit for MeHg measurements was calculated to be 0.09pmol g− 1. The accuracy of MeHg determinations was controlled by analyzing marine sediment reference materials BCR 580 and IAEA-356 (certified MeHg concentrations: 27.2 and 350pmol g− 1, respectively). Specific, potential demethylation rate constants (kd) were calculated from the equation:

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3. Results 3.1. Pore water speciation of MeHg At all sites, pore water concentrations of MeHgCl(aq), MeHgOH(aq) and MeHg+(aq) were several orders of magnitude lower than pore water concentrations of MeHgSR(aq) (organic thiol-complexes associated to dissolved organic carbon, DOC), MeHgSH(aq) and MeHgS−(aq). The sum of MeHgCl, MeHgOH and MeHg+ never comprised more than 1% of the total MeHg in pore water. The average % distribution of the species MeHgSR, MeHgSH and MeHgS− is shown in Table 1. With the exception of site Sku, the relative composition of MeHg species in pore water was similar in all surface (0–20 cm) sediments, with dominance of MeHgSH. In sub-sets that included deeper sediment layers (25–100 cm), there was a greater contribution from MeHgSR (Table 1). The reason for this is partly that the most reducing conditions (giving a predominance of MeHgSH and MeHgS−) commonly occurs at 5–20 cm depth. The relationship between the concentration of HS− in pore water and the % MeHgSR of total MeHg in pore water is shown in Fig. 1 for brackish and freshwater sediments. In principle the figure illustrates the influence of DOC on pore water MeHg speciation. In brackish sediments (Fig. 1a), having DOC concentrations between 16 and 882 mg L− 1, MeHgSR was of quantitative importance at concentrations of HS− below 20 µM, while for the freshwater sediments (Fig. 1b), having DOC concentrations between 9 and 38 mg L− 1, the organic thiols (MeHgSR complexes) were generally outcompeted by inorganic MeHg-sulfides at concentrations of HS− exceeding 0.2 µM. In Fig. 1c, the difference in pore water speciation of MeHg in brackish and freshwater sediments is highlighted for the low µM range of HS−. 3.2. Influence of additions of MeHg and Hg tracers on demethylation at two brackish water sites In order to determine the variation in demethylation rates within the two brackish water sites Sku and Köp, Me204HgCl tracer was added at a fairly constant

½Me204 Hgt2  ¼ ½Me204 Hgt0 e−kdt where [Me204Hgt2] is the concentration of Me204Hg at the end of the incubation, [Me204Hg t0] is the concentration of Me204Hg at the start of the incubation, and t is time in days. The use of the constant kd is preferred over absolute demethylation rates, since the specific rate constant is supposed to be independent of applied tracer concentration (Hintelmann et al., 2000). In one experiment the variation in demethylation rates was determined among samples taken at different depths and at different places within two brackish water sites (two depth profiles of 0–100 and 0–25 cm at site Sku and three depth profiles of 0–20, 0–25 and 0–40 cm at site Köp). In this experiment, a fairly constant concentration of isotope enriched Me204HgCl was added to the samples at each site (300– 900pmol g− 1 at Sku and 50–200pmol g− 1 at Köp). Because of substantial variations in ambient concentrations of MeHg and Hg within and among the sediment profiles, the ratio between added Me204Hg spike and ambient MeHg varied between 95 and 5800% at site Sku and between 53 and 1473% at site Köp. It was found that both the demethylation rate (ppb d− 1) and the kd (d− 1) were dependent on the added tracer expressed as % of ambient MeHg (and ambient Hg) in this experiment. In another experiment, demethylation rates were determined in surface sediments (0–10 cm) across all sites. The aim was to add Me204HgCl corresponding to 40% of ambient MeHg. However, since we used ambient concentrations of Hg as predictor, the additions of Me204HgCl fell within the range of 1–165% of ambient MeHg (20–60% for most samples). In this experiment demethylation rates (ppb d− 1) and kd (d− 1) were indicated to be non-related to the added quantity of MeHg (and Hg) tracer (expressed as % of ambient MeHg).

Fig. 3. Relationship between a) ambient MeHg (pmol g− 1) in sediments and the specific potential demethylation rate constant, kd (d− 1), b) ambient Hg in sediments and the specific potential demethylation rate constant, kd (d− 1), for all sites. Filled squares are brackish water sites (Sku and Köp) and open diamonds are freshwater sites (Mar, Tur, Sva, Nöt and Kar).

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final concentration in relation to the mass of sediment samples. Because the sediments at both sites showed large variations in ambient MeHg (and Hg) within and among profiles, the concentration of tracer in relation to ambient MeHg (and Hg) came to vary substantially (98–5800% at Sku and 53–1473% at Köp). For the two depth profiles at Sku (0–25 and 0–100 cm) there was a significant (p b 0.001), positive relationship between kd and added tracer expressed as % of ambient MeHg (Fig. 2a). Because also 201Hg(NO3)2-tracer was added to the same sample (in a parallel methylation experiment), kd was also found to be significantly (p b 0.001) correlated to added Hg expressed as % of ambient Hg (Fig. 2b). For the three depth profiles at Köp (0–20, 0–25 and 0–40 cm), on the other hand, no demethylation was detected when MeHg tracer additions exceeded 500% of ambient MeHg. As a consequence, a significant (p b 0.001) logarithmic, negative relationship was obtained between kd and added MeHg tracer in % of ambient MeHg (Fig. 2a) and between kd and added Hg tracer in % of ambient Hg (Fig. 2b). Note that the absolute concentration of added tracer showed no significant relationship with either the demethylation rate (ppb d− 1), or the kd (data not shown). Thus, as further illustrated in Fig. 2c–f, samples with low and high concentrations of MeHg and Hg at the two sites showed fundamentally different responses to MeHg (and Hg) tracer additions. At site Sku, the demethylation rate and kd were inversely related to ambient MeHg and Hg, whereas at site Köp the demethylation rate and kd were positively related to ambient MeHg and Hg. As can be seen in Figs. S1 and S2 (Supplementary information), the relationships at Sku and Köp were caused by variations in kd and ambient MeHg both with depth within a profile and between profiles. 3.3. Influence of ambient concentrations of Hg and MeHg on demethylation The variation in methylation rates were compared for surface sediments (0–10 cm) across all sites. In this experiment the addition of tracer was adjusted to yield a constant percentage of ambient MeHg (aiming at 40%) in each sample. The final range for all samples was 1–165% of ambient MeHg, with most samples within the range 20–60%. Plots revealed no significant relationships between kd and added MeHg tracer expressed as % of ambient MeHg (or added Hg tracer in % of ambient Hg), data not shown. Data on

average kd for these sites have previously been reported (Drott et al., 2008). See Table S1 (Supplementary information) for a compilation of kd values for all sampling occasions. There was a significant (p b 0.001) positive relationship between ambient concentrations of MeHg in the surface sediments and kd (Fig. 3a) across all sites. In contrast, there was no significant (p N 0.05) relationship between ambient Hg in sediments and kd, despite a wide span in ambient Hg, ranging up to a maximum of 1104 nmol g− 1 (Fig. 3b). 3.4. Influence of pore water MeHg speciation on demethylation For data from the 0–10 surface sediments, for which demethylation rates were shown to be independent of tracer additions, the possible influence of MeHg pore water speciation on kd was evaluated. For a sub-set of sites with relatively high concentrations of total Hg in sediments (sites Köp, Sku and Mar), there were no significant (p N 0.05) relationships between concentrations of MeHg species in pore water and kd, or between the total concentration of MeHg in pore water and kd. Note that for Sku, data from the area where dredged sediment had been disposed were excluded, as the properties of the sediments in this area likely had been altered by the dredging (Table 1). The average kd (± SD) at the high total Hg sites was 0.20 ± 0.24d− 1 and the average total MeHg concentration (± SD) in the sediment was 209 ± 189pmol g dw− 1. In contrast, for a sub-set of southern freshwater sites with low concentrations of total Hg in sediments (sites Tur, Sva and Nöt), a significant positive relationship was obtained between concentrations of MeHg-sulfides in pore water and kd, regardless of whether the sum of MeHgSH(aq) and MeHgS−(aq), or either species individually, were used. However, because of uncertainties about the magnitude of the pKa-value for MeHgSH, only the relationship with the sum of both species (p = 0.02) is shown (Fig. 4c). There was no significant (p N 0.05) relationship between the total concentration of MeHg in pore water and kd, nor between the concentration of MeHgSR and kd, for Tur + Sva + Nöt (Fig. 4a and b). Note, however, that the sum of MeHgSH and MeHgS− dominated the pore water speciation of MeHg. For the sub-set of samples from Tur + Sva + Nöt, the average kd (± SD) was 0.11 ± 0.07d− 1 and the average

Fig. 4. Relationship between MeHg species concentrations and specific potential demethylation rate constant, kd (d− 1), at low ambient Hg, southern freshwater sites (Tur, Sva and Nöt); a) total pore water MeHg, b) MeHgSR(aq), c) MeHgSH(aq) + MeHgS−(aq), and d) ambient MeHg in sediment.

A. Drott et al. / Marine Chemistry 112 (2008) 93–101 total sediment MeHg concentration (± SD) was 75 ± 82pmol g dw− 1. Thus, as compared to the sites with high concentrations of ambient Hg in sediments (Köp + Sku + Mar), both kd and ambient concentrations of MeHg in sediments were lower and less variable at the less contaminated sites Tur + Sva + Nöt. When the northern freshwater site Kar, also having low ambient Hg, was included with the southern freshwater sites Tur + Sva + Nöt there was no significant (p N 0.05) relationship between concentrations of MeHg species in pore water and kd. The average kd at Kar was 0.05 ± 0.06d− 1.

4. Discussion 4.1. Pore water speciation of MeHg The dominance of MeHgSH, MeHgS− and MeHgSR in pore water suggests that the solubility of MeHg in reduced sediments mainly is controlled by the concentrations of inorganic sulfides and DOC (Table 1). The solubility of MeHg per se has several implications, largely controlling migration of MeHg from the sediment to the water column, uptake and biomagnification of MeHg in the aquatic food web, and possibly also access to MeHg for demethylating microbes. Given that differences in pKa values between RSH (9.96) and H2S (7.0) largely compensate for differences in stability constants for the formation of MeHgSH and MeHgSR (log K = 14.5 and 16.5, respectively), the relative importance of RS− and HS− for the speciation of MeHg in pore water is highly dependent on the concentrations of DOC (as a proxy for RSH) and inorganic sulfides. This is illustrated in Fig.1, where high concentrations of DOC at the brackish water sites extend the range within which thiols may compete with HS−(aq) for MeHg. 4.2. Variation in demethylation within two brackish water sites The addition of high and variable concentrations of MeHg (and Hg) tracers in relation to ambient MeHg and Hg resulted in dramatically different responses at the two brackish water sites (Fig. 2a–f). At site Sku samples taken at 0–25 cm in profile 2 (Fig. S1, Supplementary information) and samples at 3–12 cm in profile 1 all showed high demethylation rates and relatively low concentrations of ambient MeHg. Because all samples were receiving the same absolute concentration of MeHg and Hg tracer, the relationships in Fig. 2a–f suggest that demethylation was stimulated by additions of MeHg (or Hg) and that samples with low ambient concentrations of MeHg (and Hg) were stimulated more than samples with high ambient MeHg (and Hg). This may further suggest that low ambient concentrations of MeHg reflect a high rate of demethylation in profile 2 and at 3–12 cm depth in profile 1 at site Sku. At site Köp it was the other way around; samples with high ambient MeHg (and Hg) generally showed the highest demethylation rates. The relationship was non-linear with a threshold at a tracer addition corresponding to 500% of ambient MeHg, above which no demethylation could be detected. Differences among samples at Köp were both related to variations among the three profiles and with depth within profile 1 (Fig. S2, Supplementary information). Considering that both the demethylation rate and kd were lower at site Köp, as compared with site Sku, we suggest that samples with low or undetectable demethylation rates at site Köp simply experienced a toxic effect by the added MeHg (or Hg). The maximum concentration of MeHg tracer applied to sediments from Sku and Köp was 210 ng Me204Hg g− 1. This

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value can be compared with Ramlal et al. (1986), who reported no change in kd with MeHg tracer additions up to 44 µg MeHg g− 1. As the naturally occurring MeHg was undetected in that study, the tracer addition obviously was many orders of magnitude greater than ambient MeHg. Marvin-DiPasquale and Oremland (1998) found both decreasing and unaffected kd with increasing tracer additions at concentrations up to 4.5 µg MeHg g− 1, in Florida Everglades (U.S.) peat sediment. Both the concentration ranges of inorganic sulfides and DOC, and the relative and absolute composition of MeHg species in pore water was fairly similar at Sku (0–100 cm) and Köp (0–40 cm, Table 1), thus there should be no major differences in availability of the added spike between these two sites. Hintelmann et al. (2000), using stable isotope tracers, concluded that there was no difference in availability for MeHg demethylation between added spike and ambient MeHg (the added spike corresponded to 10–80% of ambient MeHg). Induction of the mer operon by MeHg has not been unequivocally demonstrated. Because Hg(II) was added along with MeHg to our samples, the significant increases in the specific potential demethylation rate constant with added tracer at Sku, may suggest an induction of the reductive mermediated demethylation mechanism by either Hg(II) or MeHg. However, it is also possible that as MeHg becomes a fairly abundant C1 compound, degradation by oxidative demethylation is increased. The toxic effect at site Köp suggests that the microbial community at this site was less well adapted to high levels of MeHg and Hg(II). It can be noted that ambient levels of MeHg and Hg(II) at the two sites were fairly similar (Table S1, Supplementary information). Since biological activity was not controlled by sterilization, contribution from abiotic demethylation processes cannot be ruled out in this experiment. 4.3. Influence of ambient Hg and MeHg, and pore water speciation of MeHg, on demethylation The significant positive relationship between ambient MeHg in surface sediments and kd (Fig. 3) across all sites, and the lack of significant positive relationship between ambient Hg and kd, could be interpreted as an induction of demethylating microbes by MeHg. Previous work on induction of reductive mer-mediated demethylation by MeHg is very limited. To the best of our knowledge, the only study in which induction of reductive mer-mediated demethylation by MeHg has been tested (and supported) for a broad-spectrum resistant organism (Pseudomonas K62) is reported by Furakawa et al. (1969). The purity of the MeHgCl used for induction in their experiment is not known, and it may have contained traces of Hg(II), as suggested by Selifonova et al. (1993). Thus, at this point, induction of reductive mermediated demethylation by MeHg has not been unequivocally demonstrated. Also, it should be noted that within the sites Köp, Sku and Kar, and across all sites, there was a significant positive relationship between ambient Hg and MeHg in sediments (Drott et al., 2008). This limits the possibilities to draw strong conclusions about induction mechanisms. However, there is clearly a need for more work, to investigate whether reductive mer-mediated demethylation can be induced by MeHg, or not.

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Set into perspective to previous studies, the ambient Hg concentrations in sediments at the high contaminated sites should be sufficient for the reductive mer-operon mechanism to dominate demethylation. The average ambient Hg concentration (± SD) at the high contaminated sites Köp, Sku and Mar was 185 ± 249 nmol g− 1, to be compared with a maximum Hg concentration of 106 nmol g− 1 in freshwater sediments downstream the New Idria mercury mine (California, U.S.), where reductive mer-mediated demethylation was observed to dominate (Marvin-DiPasquale et al., 2000). In contrast, an average ambient Hg concentration of 8.2 ± 6.0 nmol g− 1 at the low contaminated sites Tur, Sva and Nöt is below the threshold at which oxidative demethylation could be expected to dominate (Marvin-DiPasquale et al., 2000). There was no marked difference in pore water MeHg species composition between surface sediments from high contaminated sites (Köp, Sku and Mar) and low contaminated sites (e.g. Tur, Sva and Nöt, Table 1). Thus, the difference in relationships between pore water MeHg species and kd obtained for these two sub-sets of sites cannot be explained by differences in pore water MeHg species composition. The lack of a relationship between concentrations of MeHg species in the sediment pore water and kd for the high contaminated sites Köp, Sku and Mar may suggest that demethylation is not limited by the availability of MeHg species at these sites. Zero-order kinetics are expected for active or carrier-mediated transport at high concentrations of the compound being taken up (e.g. Michaelis–Menten kinetics). However, active transport of MeHg compounds during reductive mer-mediated demethylation has not been demonstrated [e.g. Barkay et al., 2003]. Thus, it is more likely that another factor/-s limit demethylation rates at the high ambient Hg sites. In contrast, the significant positive relationship between concentrations of MeHgSH, and MeHgSH + MeHgS−, and kd for the sub-set Tur + Sva + Nöt suggests an influence of pore water MeHg speciation on demethylation. If an oxidative demethylation mechanism is assumed (because of lower ambient Hg), the relationship with the concentration of MeHgSH(aq) may suggest a passive uptake mechanism. However, at low concentration of the compound being taken up, there may also be a significant positive relationship if active or carrier-mediated transport is taking place. The average concentration (± SD) of MeHgSH(aq) + MeHgS−(aq) was 9.4 ± 7.0pM at the low contaminated sites Tur, Sva and Nöt and 53 ± 99pM at the high contaminated sites Köp, Sku and Mar. At present, the knowledge about oxidative demethylation is limited and it is not certain in what form and by what mechanism MeHg is taken up. It can be noted that the pKa-value for MeHgSH was estimated from statistical relationships (Dyrssen and Wedborg, 1991) and has not been confirmed experimentally. Thus, the value of 7.5 is subjected to substantial uncertainty and if the pKa for MeHgSH is in fact greater, the contribution from MeHgS− is smaller than calculated. Therefore, the significant positive relationship between MeHg-sulfides and kd may be explained by a passive uptake of the uncharged MeHgSH species. The negative relationship between MeHgSR and kd (Fig. 4b) for the southern freshwater sites indicates that MeHgSR is largely unavailable to demethylating microbes. This is in line with Marvin-DiPasquale et al. (2000), who suggested that the formation of MeHg–organic complexes

was a possible cause of observed low kd in organic-rich systems (Everglades, Florida, U.S.), where oxidative demethylation was dominant. The reductive demethylation pathway involving sulfate reducers (Baldi et al., 1993) does not seem to contribute significantly to demethylation in the sediments of the present study, as there were no significant (p N 0.05) positive relationships between pore water HS− (or sulfate) and kd for any site or sub-set of sites (data not shown). The possible contribution from abiotic demethylation mechanisms cannot be ruled out, as no sterilized control was included. Abiotic demethylation mechanisms can be expected to contribute most significantly at large sediment depths, where microbes are poorly supported with electron donors (organic matter) from the water column. Given also that all work was done in sediments that were kept in darkness, it should be safe to assume that there is no contribution from photodegradation in the present study. 5. Conclusions Equilibrium speciation calculations showed that inorganic MeHg-sulfides and bisulfides, and organic MeHg-thiols, were the dominant MeHg species in the sediment pore water at all sites. When MeHg and Hg tracer additions were applied, the specific potential demethylation rate constant kd (day− 1) was influenced by the amount of tracer added in relation to ambient MeHg and Hg. Additions of MeHg and Hg tracers to sediment samples from depth profiles (0–100 cm) resulted in enhanced demethylation rates with MeHg and Hg tracer additions (expressed as % of ambient MeHg and Hg) at one brackish water site. Similar relative additions to depth profiles (0–40 cm) at another brackish water site resulted in toxic effects and non-detectable demethylation above a threshold at MeHg tracer addition corresponding to approximately 500% of ambient MeHg. Thus, for apparently similar types of brackish water sediments striking differences were observed. In surface sediments (0–10 cm) across all seven sites, significant correlations between kd and ambient MeHg (but not ambient Hg) may suggest that demethylation is enhanced in samples with an enhanced net production of MeHg. In a sub-set of freshwater surface (0–10 cm) sediments with lower ambient concentrations of total Hg, kd was positively correlated to concentrations of MeHgSH, and the sum of MeHgSH and MeHgS−. We interpret this as a possible passive uptake of inorganic MeHgSH during oxidative demethylation. In contrast, there was no relationship between pore water MeHg speciation and kd at sites with higher ambient Hg concentrations. Acknowledgments Bengt Andersson and Tom Larsson are greatly acknowledged for assistance in the laboratory. The authors are grateful to Umeå Marine Sciences Centre and the Department of Ecology and Environmental Sciences at Umeå University for providing sampling equipment. This work was financed by the Centre for Environmental Research, Umeå, the North Sweden Soil Remediation Centre (MCN) — European Union Structural Funds and New Objective 1 (contract No. 11312534-00), Swedish EPA (contract E-38-04), by the Knut and Alice Wallenberg Foundation and by the Kempe Foundation.

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