Assessing the Stability of Mercury and Methylmercury in a Varved

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Environ. Sci. Technol. 2008, 42, 4391–4396

Assessing the Stability of Mercury and Methylmercury in a Varved Lake Sediment Deposit JOHAN RYDBERG,* VERONIKA GÄLMAN, INGEMAR RENBERG, AND RICHARD BINDLER Environmental Change Assessment Group, Department of Ecology and Environmental Science, Umeå University, SE-901 87 Umeå, Sweden LARS LAMBERTSSON Department of Chemistry, Umeå University, SE-901 87 Umeå, Sweden, A N T O N I O M A R T ´I N E Z - C O R T I Z A S Departmento de Edafologia y Quimica Agricola, Faculty of Biology, University of Santiago de Compostela, E-15706 Santiago de Compostela, Spain

Received December 20, 2007. Revised manuscript received March 6, 2008. Accepted March 12, 2008.

Using lake sediments to infer past total mercury and methylmercury loading to the environment requires that diagenetic processes within the sediment do not significantly affect the concentrations or net accumulation rates of the mercury species. Because carbon is lost during early sediment diagenesis, the close link between carbon and mercury raises the question of how reliable lake sediments are as archives of total mercury and methylmercury loading. In this study we used a series of freeze cores taken in a lake with varved (annually laminated) sediment to assess the stability of total mercury and methylmercury over time. By tracking material deposited in specific years in cores collected in different years, we found that despite a 20–25% loss of carbon in the first 10–15 years, there was no apparent loss of total mercury over time; hence, lake sediments can be considered as reliable archives. However, over the first 5–8 years after sedimentation, about 30–40% of the methylmercury was lost (a decrease of 0.025–0.030 µg MeHg m-2 yr-1), suggesting that sediment profiles showing increasing methylmercury concentrations toward the sediment surface are in large part an artifact of diagenetic processes (net demethylation), rather than a record of changes in methylmercury loading.

Introduction An important premise for reconstructing environmental changes using lake sediments is that the environmental parameter or signal of interest is preserved in the sediment. After deposition, the sedimented material is affected by a number of diagenetic processes, both biotic and abiotic, which may alter its properties (diagenesis meaning chemical and physical changes that occur within the sediment after deposition). Loss of organic matter is one of the effects of * Corresponding author phone: +46(0)90-7867947; fax: +46(0)907866705; e-mail: [email protected]. 10.1021/es7031955 CCC: $40.75

Published on Web 05/09/2008

 2008 American Chemical Society

diagenesis (1, 2), and this has possible implications for a large number of elements in the sediment, especially those associated with organic matter, such as mercury. Diagenesis might also change the redox conditions that affect the preservation of different elements, such as iron and sulfur (3, 4), hence possibly also mercury. Lake sediments are often used as an archive to study past variations in mercury loading to the environment (5–8), and sediments are a significant component in lake biogeochemical models of mercury cycling (9). Thus, it is of importance to know whether mercury is lost from the sediment together with carbon, relocated within the sediment, as is the case for cesium (10), or whether it is unaffected by diagenesis. For example, Gobeil et al. (11) and Beldowski et al. (12) have suggested that mercury may be mobile in some marine environments with low sedimentation rates. Even though a large number of studies have examined mercury in lake sediments, there are few published investigations actually dealing with the effects of diagenetic processes. One of the few examples is that of Lockhart et al. (13), who studied sediment deposits from three lakes polluted by point sources. In one of the lakes they compared total mercury concentration profiles from three sediment cores collected in different years (1971, 1978 and 1995), whereas in the other two lakes they compared the total mercury concentration profiles in dated sediment cores with the known pollution histories of the sites. Even if their research indicates that sediment total mercury profiles are reliable as archives for heavily polluted sites, the large amount of mercury emitted to the studied lakes and the uncertainties of the 210Pb-dating made it impossible to assess whether small-scale migration (small amounts of mercury over distances up to centimeters) or loss of mercury from the sediment had occurred. The possibility of small-scale migration has implications when using sediment deposits to study the history of mercury loading to lakes, especially in studies with a high temporal resolution or at sites where small changes in concentration may be of interest. Knowledge about migration of mercury in lake sediments is also important for the understanding of mercury cycling in lake ecosystems because a substantial amount of the mercury present in lakes occurs in the upper centimeters of the sediment (14). Most of the mercury in lake sediment is in inorganic form, but a small fraction is present as methylmercury. Methylmercury is highly toxic, readily bioaccumulated, and is considered as the mercury species of greatest importance for mercury levels in piscivorus fish and fish-eating wildlife (15). In lake sediments, methylmercury concentrations typically increase toward the sediment surface (16–18), which raises the question of whether this increase represents an actual increased loading of methylmercury to the environment or whether it is the result of diagenetic processes in the sediment. Some studies have pointed out that the behavior of methylmercury is highly dependent on the conditions within the sediment, such as redox properties, temperature, and sediment composition; hence, its fate can be very different from inorganic mercury species due to, for example, diffusive loss to the water column and methylation/demethylation processes (19–21). In this paper we address the question of how diagenesis affects total mercury and, based on a smaller number of samples, methylmercury concentrations in a lake sediment deposit. For this purpose we use a unique sample collection consisting of eight freeze cores collected in different years, from 1979 until 2007, in lake Nylandssjön, which has varved VOL. 42, NO. 12, 2008 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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TABLE 1. Results of Reference Samples Analyzed to Evaluate the Accuracy of the Total Mercury Analyses g-1)

certified value (ng number of samples (n) result (ng g-1)

CANMET, LKSD-4 sediment

low ash peat (27)

190 1 193

169 ( 7 (recommended value) 21 162 ( 6

(annually laminated) sediment. With this sample collection we are able to track the sediment material deposited in a specific year and study how the mercury and methylmercury concentrations change over time. For example, we can track changes in mercury concentration in the sediment material accumulated in 1978, (the newly deposited material in the 1979-core), after 6 years in the 1985-core, and up to 25-years in the 2004-core. From a previous study in Nylandssjön, done on the same samples as those of the present study, we know that between 20 and 25% of the carbon accumulated in the surface varve is lost during the first 10–15 years following accumulation (2), making it possible to study whether this loss of carbon also affects mercury concentrations.

Materials and Methods Study Site. Nylandssjön is a small (0.28 km2) dimictic lake situated in northern Sweden (62°57′ N, 18°17′ E, 34 m above sea level). It has two deep basins, the deeper is 17.5 m and the shallower is 14.3 m. Following lake isolation from the Gulf of Bothnia about 3400 years ago due to land-upheaval, fairly good varves were preserved in the sediment (22). Over time the quality of the varves deteriorated and varve sequences occur only sporadically. For the last ∼80 years, once again, very distinct varves are preserved in the sediments of both lake basins. This resumption of varve formation in the early 1900s, as well as the relatively high sediment accumulation rate, is due to cultural eutrophication (23, 24). For further details about the lake see Petterson et al. (25) and Gälman et al. (24). Sampling. All sediment cores were collected in Nylandssjön during late winter using a freeze corer (26). The eight sediment cores from the 17.5 m basin were collected from 1979 to 2007 (1979, 1985, 1989, 1993, 1997, 2002, 2004, and 2007) and the core from the 14.3 m basin was collected in 2004. After collection, the cores were stored in a freezer room (-18 °C) until subsampling, which was done in 2004–2007. For subsampling, the frozen cores were first cleaned using a woodworker’s plane, and then each varve was individually cut out using a scalpel. The upper part of the distinct black winter layer, which marks the transition from winter to spring, was used as the divider between years. At the sediment-water interface the varves are about 1 cm thick, compaction makes the varves thinner with time, and after 50 years they are typically about 2–3 mm. Each varve sample represented 7–10 cm2 of the lake bottom area and gave a dry sample mass of 150–300 mg. Typically, the 35 most recent varves were cut from each core; varves older than 35 years are thin due to compaction and are therefore difficult to cut at annual resolution. In total, the overlapping sequence of freeze cores contains 276 samples spanning the period 1948–2007. In 2006 we also sampled deeper sediments using a Russian peat corer. These cores were subsampled every tenth centimeter from the resumption of varve formation (∼1920s) and down to ∼155 cm sediment depth, which corresponds to the isolation horizon in the sediment (∼3400 BP). The age of the deeper sediment samples (prior to 1948) was estimated by extrapolating varve thickness from periods with visible varves to the entire core sequence. Analyses. All samples were freeze-dried and homogenized prior to analyses. Carbon was analyzed using a Perkin-Elmer 2400 Series II CHNS/O-analyzer (operated in CHN-mode), 4392

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NCS ZC73002 - soil NIST 1515 - apple leaves 60 ( 9 15 55 ( 6

44 ( 4 7 45 ( 2

and these data have been published in Gälman et al. (2). For total mercury analyses, a Leco AMA 254-analyzer was used, which requires no further sample preparation. The calibration curve for total mercury analyses was based on analyses of different masses (10–200 mg) of MESS-3 (National Resources Canada, Institute for National Measurement Standards; certified value 91 ( 9 ng Hg g-1). Analysis quality was controlled using standard reference materials (CANMET LKSD-4, low ash peat (27), NCS ZC73002-Soil, NIST 1515Apple leaves). All reference samples were within their certified range (Table 1), and replicate analyses of sediment samples were within ( 5% (n ) 30). Methylmercury was analyzed in 2007 according to Lambertsson et al. (28). In brief, prior to sample treatment, sediment samples were spiked with an aqueous solution of Me200Hg+ to 0.5% of the total mercury concentration as a species-specific isotope standard for isotope dilution analysis and was equilibrated overnight at 23 °C. Methylmercury was then solid–liquid extracted from the samples using a mixture of KBr/CuSO4/H2SO4/CH2Cl2, derivatized with NaB(C2H5)4 (28) and analyzed by GC-ICP-MS (Agilent 6890N GC, 7500a ICP-MS). The method precision for methylmercury determinations was 5% RSD, based on replicate sample analyses (n ) 3), and the method detection limit was 0.03 ng MeHg g-1, calculated as 3-times the standard deviation (3σ) of the blank. Sulfur was analyzed with an energy dispersive XRFanalyzer for light elements (Al, Cl, Mg, P, S, Si) (29). Accuracy and precision were assessed using standard reference materials, which were within ( 10% of certified values.

Results There is a good agreement in total mercury concentration profiles for cores collected in different years from the 17.5 m basin, even though 28 years have elapsed between the collection of the first and the last core (Figure 1). Total mercury concentrations in sediment from the 17.5 m and the 14.3 m basins also show good agreement (Figure 2a), and the net annual total mercury accumulation rates display the same trend even though the absolute levels are somewhat lower in the 14.3 m basin due to lower total sediment accumulation rate (Figure 2b). When we track the sediment material from a specific year (varve), that is, from when it was newly deposited material (a surface varve) and onward in subsequent cores as that year’s sediment material gets older, we find that total mercury concentrations are almost constant through time (Figure 3a). Methylmercury concentrations, on the other hand, clearly decline as the sediment material ages (Figure 3b). In the first year following sedimentation (in a surface varve) the concentration is 1.5–2 ng MeHg g-1 dry sediment (about 2% of the total mercury concentration) and after 5–8 years the concentration has decreased by 30–40% to ∼1 ng MeHg g-1. After 22 years the original methylmercury concentration of 1.5 ng MeHg g-1 in the 1984 varve has declined by ∼60% to 0.6 ng MeHg g-1. In sediment deposited prior to ∼1900 total mercury concentration fluctuates between 20 and 75 ng Hg g-1 dry sediment (mean of ∼50 ng Hg g-1; Figure 2a). In 1948 (the first sample with annual resolution) the concentration was ∼70 ng Hg g-1, and over the following two decades the

FIGURE 1. Total mercury concentrations (ng g-1 dry sediment) for all analyzed cores from the 17.5 m basin, vertical lines represent the time periods covered by the different cores. concentration increased and reached its highest value (∼100 ng Hg g-1) in 1968. For the last 35 years the total mercury concentration has been more-or-less stable, varying around a mean of ∼90 ng Hg g-1. However, net annual total mercury accumulation rates (Figure 2b) show a long-term declining trend since the late 1960s. The net annual accumulation rate of total mercury during the late 1940s was ∼2.3 µg Hg m-2 yr-1, which increased in the late 1960s to ∼3.5 µg Hg m-2 yr-1 and has since declined to a mean value of ∼2.1 µg Hg m-2 yr-1 for the last 10 years. The concentration of sulfur, which also can be important for mercury retention in the environment (30), seems to be unaffected in the sediment of this lake over time (Figure 3c). Recent work, done on a small number of samples, indicates that between 30 and 60% of the sulfur in the sediment in Nylandssjön is present in the form of thiols (31), which is the most important sulfur form for the binding of mercury to organic matter.

Discussion Gray-scale image analysis (25) showed that, apart for compaction, the varve structure is preserved over time. When Gälman et al. (24) compared sediment cores from different locations within Nylandssjön’s deep-basins they also found that the carbon content is spatially consistent, both within and between basins. However, recent analyses of the carbon content of the same samples as in this mercury study show that diagenetic changes do occur within the sediment (ref 2; Figure 3d). During the first 20 years after sediment accumulation, about 20–25% of the carbon is lost due to diagenetic processes. Despite this loss of carbon the total mercury record of all eight sediment cores shows very similar large-scale patterns (Figure 1). For example, there are consistent, well-defined concentration peaks in the early 1950s and later 1960s and troughs in the late 1950s, mid1970s, and in 1996. This clearly illustrates that the sediment is a good archive over past total mercury loading to the lake

basin, and this is further strengthened by the very good agreement between the records from the two lake basins (Figure 2a). The small differences that do exist between different cores can largely be explained by two factors. The primary factor is that the sediment surface on the lake bottom in the 17.5-m deep basin is not completely even, and hence, we expect some slight spatial differences in the relative mixture of heavy material (mineral grains) and light material (organic material) in the different cores. The second factor is that there may be small differences between samples as a result of imperfect separation between varves during subsampling. The comparison of total mercury concentrations of specific years in all eight cores reveals that no significant changes occurred during the ∼25 years of sediment aging in which we can track the samples (Figure 3a). There is a small tendency for the total mercury concentration of a given year to increase over time, but this is not statistically significant because the change is within the analytical error ((5%). A small increase would be expected given that the diagenetic loss of carbon corresponds to a loss of ∼4% of the total sediment dry mass over 10–20 years (∼20% loss of the initial carbon concentration, which was ∼20% of the dry sediment mass). The years 1996 and 2001 have a larger difference in total mercury concentration between the different cores than other years, up to 30%. In these years there is a larger than average difference in total sediment accumulation rate (µg m-2 yr-1) between the different cores. However, even though the concentration varies among cores, the total mercury accumulation rates are very similar for all samples from these years. From the mercury data itself we cannot tell whether the mercury in a specific varve is first released as the organic matter is degraded and then simply rebound to other substances in the sediment or if the mercury is attached to a fraction of the organic matter that is unaffected by diagenesis. However, changes in C/N ratio and the behavior of sulfur in the sediment give information that can help in the interpretation. Gälman et al. (2) have found that the C/N ratio in Nylandssjön sediment increases with time. In newly accumulated material (deposited during the previous year) the C/N ratio is 9.5–10.5 (molar ratio), but it increases to 11.5–12.5 during the first 15–20 years. A likely explanation for this increase in the C/N ratio is a preferential mineralization of algal material, which has C/N ratios of about 5–11 (1), and a preferential preservation of terrestrially derived organic matter, which generally has higher C/N ratios. The terrestrially derived organic matter has already been subjected to degradation in the terrestrial environment and is therefore relatively resilient to further degradation in the sediment. The total sulfur concentration in sediment varves from Nylandssjön remain unchanged with time (Figure 3c), and much of the sulfur is present as thiols (31). Reduced sulfur groups on organic matter (thiols) have been found to be of great importance for the retention of mercury in boreal forest soils (32–34), and because a substantial part of both the organic and inorganic material of most lake sediments is derived from the catchment, we can argue that this link between sulfur and mercury is also important in lake sediments. On the basis of basic knowledge about sediment formation (35), the link between sulfur and mercury (33, 34, 36), the increasing C/N ratios, and the unchanged sulfur concentrations in Nylandssjön’s sediment, we suggest that the mercury that is accumulated in Nylandssjön’s sediment largely is bound to reduced sulfur groups attached to old terrestrially derived organic matter and that this fraction of the organic matter is not, to any larger extent, subjected to degradation in the sediment. In contrast to total mercury, methylmercury concentrations decrease by 30–40% in the first 5–8 years and by ∼60% VOL. 42, NO. 12, 2008 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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FIGURE 2. (a) Mean total mercury concentrations (ng g-1 dry sediment, calculated using all eight cores) for the last 60 years as well as samples from deeper sediments (back to 1000 BC) in Nylandssjön. Data from the 17.5 m basin are shown as filled circles (with error bars) and those from the 14.3 m basin as open circles. (b) Mean net annual total mercury accumulation rates (AR) since 1948. The star (44) and the open squares with error bars (www.ivl.se) show mercury concentrations (µg g-1 dry mass) in mosses from northern Sweden. (c) Mean carbon-normalized total mercury levels. over 20 years (Figure 3b). The initial 5–8 year decline in methylmercury corresponds to a decrease of 0.025–0.030 µg MeHg m-2 yr-1. This decline in the methylmercury concentrations may be due to either a change in the ratio between methylation- and demethylation rate with sediment depth, mineralization of deposited methylmercury, or a transport of methylmercury from the sediment to the water column. Because the decline in methylmercury concentration of ∼1 ng MeHg g-1 is within the analytical error ((5%) of the total mercury analyses, we cannot determine whether methylmercury remains in the sediment in an inorganic form or is transported to the water column. However, our results clearly show that lake sediments are not an accurate archive of past methylmercury levels. Thus, the decline in methylmercury concentrations with depth in sediment cores found in a number of lakes (16–18) most likely is a result of diagenesis and does not reflect any recent increase in the input of methylmercury to the sediment. Even if the general view is that methylmercury in lake sediments mainly is the result of in situ methylation (21), we have data that allow us to speculate on the origin of the methylmercury in Nylandssjön (in situ methylation vs deposition). We know from Eckley et al. (37) that methylmercury can be present in the hypolimnion of lakes similar to Nylandssjön (with the oxic/anoxic boundary in the water column rather than in the sediment). From Shchukarev et al. (31) we also know that oxidized species of sulfur (SO42-) and iron (different forms of FeOOH) exist in deeper sediment layers (∼20 years old) in Nylandssjön, possibly indicating a low activity of sulfur- and iron-reducing bacteria, which have 4394

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been reported to be responsible for methylation of mercury in lake sediments (21, 38). On the basis of this presence of oxidized sulfur and iron species in deeper sediment layers and the possibility of the water column (37), sediments in the littoral zone (20), and the catchment (39, 40) to act as a source for methylmercury, we suggest that at least part of the methylmercury in Nylandssjön’s sediment is deposited rather than produced in the sediment by sulfur- and ironreducing bacteria. However, to determine whether methylmercury is deposited on the sediment, or if changes in the methylation/demethylation rate ratio is the cause of the declining methylmercury trend, would require further analyses of methylation and demethylation rates in the sediment and methylmercury concentrations in the water column. Mercury emissions in Sweden and Europe were highest in the late 1960s, and since then emissions and deposition have decreased markedly (41); this pattern is reflected in a number of sediment profiles from south Swedish lakes (42). In northern Sweden, deposition was never as high as in the south; consequently, the decreasing trend in lake sediments is less pronounced (43). The total mercury concentration in the sediment of Nylandssjön shows little change over the last decades, with total mercury concentrations around 90 ng Hg g-1. However, a significant change in the sediment material composition in the late 1980s toward higher carbon concentrations (2, 24) makes it more appropriate (apart from the most recent years where there still is considerable carbon loss) to look at carbon-normalized mercury levels (Hg/C ratios, by weight) or net annual total mercury accumulation rates rather than total mercury concentrations (Figure 2,

sediment mixing, which is not the case for most lake sediments; therefore, it could be argued that results from varved sediments may not be applicable for all lakes. Nevertheless, our findings from Nylandssjön’s sediment agree well with those of other studies done on nonvarved lake sediments (6, 13, 17), and it is therefore reasonable to assume that our results showing the stability of total mercury concentrations and the loss of methylmercury concentrations over ∼25 years are also applicable for nonvarved lake sediments.

Acknowledgments The XRF measurements were performed at the RIAIDT of the USC (University of Santiago de Compostela, Spain). We want to thank Manuela Costa Casais, who made the XRF measurments. Financial support was provided from the Faculty of Science and Technology at Umeå University and The Swedish Research Council.

Literature Cited

FIGURE 3. Concentration changes for total mercury (a), methylmercury (b), sulfur (c), and carbon (d; ref (2)) over time. Age 0 is when the material is newly deposited (i.e., a surface varve). panels b and c). Both the carbon-normalized mercury levels and the net annual accumulation rates of total mercury show a clear decrease since the peak in the late 1960s. In 1992 (mean age of the analyzed material is older than 10-years, i.e., the main part of carbon degradation has occurred) the carbon-normalized mercury level was ∼60% less than the peak value in 1968 but was ∼30% above pre-1800 levels. For the net annual accumulation rate of total mercury, the decrease since the late 1960s is from ∼3.5 µg Hg m-2 yr-1 to ∼2.1 µg Hg m-2 yr-1. This 40% decline is not as large as that indicated by environmental monitoring based on mosses (Hylocomium splendens and Pleurozium schreberi), which are used as a proxy for atmospheric mercury deposition (44). Between 1968/70 and 2005, mercury concentrations in mosses decreased from ∼70 to ∼10 ng Hg g-1 in the region around Nylandssjön (ref 44, www.ivl.se; Figure 2c). The larger decrease in the total mercury concentration in mosses compared to the lake sediments agrees well with the suggestion that a substantial amount of the mercury in the lake sediment is of terrestrial origin. Unlike the mosses, which only record direct atmospheric deposition, the lake sediment also receives mercury from the catchment area as well as through resuspension and transport of sediment from shallower to deeper areas, which together introduce a lag between atmospheric deposition and the lake sediment record. This study was done using varved lake sediments, which is a valuable archive for studies seeking annual resolution. An important precondition for varve formation is the lack of

(1) Meyers, P. A.; Ishiwatari, R. Lacustrine organic geochemistry — An overview of indicators of organic-matter sources and diagenesis in lake-sediments. Org. Geochem. 1993, 20, 867–900. (2) Gälman, V.; Rydberg, J.; Sjöstedtde-Luna, S.; Bindler, R.; Renberg, I. Carbon and nitrogen loss rates during ageing of lake sediment: Changes over 27 years studied in a varved lake sediment. Limnol. Oceanogr. 2008, 53 (3), 1076–1082. (3) VanCappellen, P.; Wang, Y. F. Cycling of iron and manganese in surface sediments: A general theory for the coupled transport and reaction of carbon, oxygen, nitrogen, sulfur, iron, and manganese. Am. J. Sci. 1996, 296, 197–243. (4) Holmer, M.; Storkholm, P. Sulphate reduction and sulphur cycling in lake sediments: a review. Freshwater Biol. 2001, 46, 431–451. (5) Swain, E. B.; Engstrom, D. R.; Brigham, M. E.; Henning, T. A.; Brezonik, P. L. Increasing rates of atmospheric mercury deposition in midcontinental North-America. Science 1992, 257, 784– 787. (6) Fitzgerald, W. F.; Engstrom, D. R.; Mason, R. P.; Nater, E. A. The case for atmospheric mercury contamination in remote areas. Environ. Sci. Technol. 1998, 32, 1–7. (7) Bindler, R.; Renberg, I.; Appleby, P. G.; Anderson, N. J.; Rose, N. L. Mercury accumulation rates and spatial patterns in lake sediments from west Greenland: A coast to ice margin transect. Environ. Sci. Technol. 2001, 35, 1736–1741. (8) Lamborg, C. H.; Fitzgerald, W. F.; Damman, A. W. H.; Benoit, J. M.; Balcom, P. H.; Engstrom, D. R. Modern and historic atmospheric mercury fluxes in both hemispheres: Global and regional mercury cycling implications. Global Biogeochem. Cycles 2002, 16. (9) Fitzgerald, W. F.; Engstrom, D. R.; Lamborg, C. H.; Tseng, C. M.; Balcom, P. H.; Hammerschmidt, C. R. Modern and historic atmospheric mercury fluxes in northern Alaska: Global sources and arctic depletion. Environ. Sci. Technol. 2005, 39, 557–568. (10) Smith, J. T.; Comans, R. N. J.; Ireland, D. G.; Nolan, L.; Hilton, J. Experimental and in situ study of radiocaesium transfer across the sediment-water interface and mobility in lake sediments. Appl. Geochem. 2000, 15, 833–848. (11) Gobeil, C.; Macdonald, R. W.; Smith, J. N. Mercury profiles in sediments of the Arctic Ocean basins. Environ. Sci. Technol. 1999, 33, 4194–4198. (12) Beldowski, J.; Pempkowiak, J. Horizontal and vertical variabilities of mercury concentration and speciation in sediments of the Gdansk Basin, Southern Baltic Sea. Chemosphere. 2003, 52, 645– 654. (13) Lockhart, W. L.; Macdonald, R. W.; Outridge, P. M.; Wilkinson, P.; DeLaronde, J. B.; Rudd, J. W. M. Tests of the fidelity of lake sediment core records of mercury deposition to known histories of mercury contamination. Sci. Total Environ. 2000, 260, 171– 180. (14) Meili, M.; Mercury in lakes and rivers. In Metal Ions in Biological Systems; Marcel Dekker: New York, 1997; Vol 34, 21-51. (15) Ullrich, S. M.; Tanton, T. W.; Abdrashitova, S. A. Mercury in the aquatic environment: A review of factors affecting methylation. Crit. Rev. Environ. Sci. Technol. 2001, 31, 241–293. (16) Munthe, J.; Hultberg, H.; Lee, Y. H.; Parkman, H.; Iverfeldt, Å.; Renberg, I. Trends of mercury and methylmercury in deposition, VOL. 42, NO. 12, 2008 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

9

4395

(17)

(18)

(19)

(20)

(21)

(22)

(23)

(24)

(25)

(26)

(27)

(28)

(29)

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runoff water and sediments in relation to experimental manipulations and acidification. Water Air Soil Pollut. 1995, 85, 743–748. Hines, N. A.; Brezonik, P. L.; Engstrom, D. R. Sediment and porewater profiles and fluxes of mercury and methylmercury in a small seepage lake in northern Minnesota. Environ. Sci. Technol. 2004, 38, 6610–6617. He, T. R.; Lu, J.; Yang, F.; Feng, X. B. Horizontal and vertical variability of mercury species in pore water and sediments in small lakes in Ontario. Sci. Total Environ. 2007, 386, 53–64. Matilainen, T.; Verta, M.; Niemi, M.; Uusirauva, A. Specific rates of net methylmercury production in lake-sediments. Water Air Soil Pollut. 1991, 56, 595–605. Ramlal, P. S.; Kelly, C. A.; Rudd, J. W. M.; Furutani, A. Sites of methyl mercury production in remote Canadian shield lakes. Can. J. Fish. Aquat. Sci. 1993, 50, 972–979. Benoit, J. M.; Gilmour, C. C.;. Heyes, A.;. Mason, R. P.;. Miller, C. L.;. Geochemical and biological controls over methylmercury production and degradation in aquatic ecosystems. In Biogeochemistry of Environmentally Important Trace Elements; ACS Symposium Series No. 835, 2003; pp 262-297. Renberg, I. Formation, structure and visual apperance of ironrich, varved lake sediments. Verh. Int. Ver. Theor. Angew. Limnol. 1981, 21, 94–101. Renberg, I. Photographic demonstration of the annual nature of a varve type common in N Swedish lake-sediments. Hydrobiologia 1986, 140, 93–95. Gälman, V.; Petterson, G.; Renberg, I. A comparison of sediment varves (1950–2003 AD) in two adjacent lakes in northern Sweden. J. Paleolimnol. 2006, 35, 837–853. Petterson, G.; Renberg, I.; Geladi, P.; Lindberg, A.; Lindgren, F. Spatial uniformity of sediment accumulation in varved lake sediments in northern Sweden. J. Paleolimnol. 1993, 9, 195– 208. Renberg, I. Improved methods for sampling, photographing and varve-counting of varved lake sediments. Boreas 1981, 10, 255–258. Yafa, C.; Farmer, J. G.; Graham, M. C.; Bacon, J. R.; Barbante, C.; Cairns, W. R. L.; Bindler, R.; Renberg, I.; Cheburkin, A. K.; Emons, H.; Handley, M. J.; Norton, S. A.; Krachler, M.; Shotyk, W.; Li, X. D.; Martinez Cortizas, A.; Pulford, I. D.; MacIver, V.; Schweyer, J.; Steinnes, E.; Sjobakk, T. E.; Weiss, D.; Dolgopolova, A.; Kylander, M. Development of an ombrotrophic peat bog (low ash) reference material for the determination of elemental concentrations. J. Environ. Monit. 2004, 6, 493–501. Lambertsson, L.; Lundberg, E.; Nilsson, M.; Frech, W. Applications of enriched stable isotope tracers in combination with isotope dilution GC-ICP-MS to study mercury species transformation in sea sediments during in situ ethylation and determination. J. Anal. At. Spectrom. 2001, 16, 1296–1301. Cheburkin, A. K.; Shotyk, W. Energy-dispersive XRF spectrometer for Ti determination (TITAN). X-Ray Spectrom. 2005, 34, 69–72.

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(30) Skyllberg, U.; Xia, K.; Bloom, P. R.; Nater, E. A.; Bleam, W. F. Binding of mercury(II) to reduced sulfur in soil organic matter along upland-peat soil transects. J. Environ. Qual. 2000, 29, 855–865. (31) Shchukarev, A.; Gälman, V.; Rydberg, J.; Sjöberg, S.; Renberg, I. Speciation of iron and sulphur in seasonal layers of varved lake sediments: An XPS study. Surf. Interface Anal. 2008, 40 (3-4), 354–357. (32) Haitzer, M.; Aiken, G. R.; Ryan, J. N. Binding of mercury(II) to dissolved organic matter: The role of the mercury-to-DOM concentration ratio. Environ. Sci. Technol. 2002, 36, 3564–3570. (33) Ravichandran, M. Interactions between mercury and dissolved organic matter — A review. Chemosphere 2004, 55, 319–331. (34) Skyllberg, U.; Bloom, P. R.; Qian, J.; Lin, C. M.; Bleam, W. F. Complexation of mercury(II) in soil organic matter: EXAFS evidence for linear two-coordination with reduced sulfur groups. Environ. Sci. Technol. 2006, 40, 4174–4180. (35) Håkanson, L.; Jansson, M. Principals of Lake Sedimentology; Springer-Verlag: Berlin, 1983. (36) Haitzer, M.; Aiken, G. R.; Ryan, J. N. Binding of mercury(II) to aquatic humic substances: Influence of pH and source of humic substances. Environ. Sci. Technol. 2003, 37, 2436–2441. (37) Eckley, C. S.; Watras, C. J.; Hintelmann, H.; Morrison, K.; Kent, A. D.; Regnell, O. Mercury methylation in the hypolimnetic waters of lakes with and without connection to wetlands in northern Wisconsin. Can. J. Fish. Aquat. Sci. 2005, 62, 400–411. (38) Kerin, E. J.; Gilmour, C. C.; Roden, E.; Suzuki, M. T.; Coates, J. D.; Mason, R. P. Mercury methylation by dissimilatory ironreducing bacteria. Appl. Environ. Microbiol. 2006, 72, 7919– 7921. (39) Skyllberg, U.; Qian, J.; Frech, W.; Xia, K.; Bleam, W. F. Distribution of mercury, methyl mercury and organic sulphur species in soil, soil solution and stream of a boreal forest catchment. Biogeochemistry 2003, 64, 53–76. (40) Garcia, E.; Carignan, R.; Lean, D. R. S. Seasonal and inter-annual variations in methyl mercury concentrations in zooplankton from boreal lakes impacted by deforestation or natural forest fires. Environ. Monit. Assess. 2007, 131, 1–11. (41) Lindqvist, O.; Johansson, K.; Aastrup, M.; Andersson, A.; Bringmark, L.; HovseniusG.; Håkanson, L.; Iverfeldt, Å.; Meili, M.; Timm, B. Mercury in the Swedish environment - recent research on causes, consequences and corrective methods. Water Air Soil Pollut. 1991, 55, xi–261. (42) Bindler, R.; Olofsson, I.; Renberg, I.; Frech, W. Temporal trends in mercury accumulation in lake sediments in Sweden. Water Air Soil Pollut.: Focus 2001, 1, 343–355. (43) Lindeberg, C.; Bindler, R.; Bigler, C.; Rosén, P.; Renberg, I. Mercury pollution trends in sub-arctic lakes in the northern Swedish mountains. Ambio 2007, 36, 401–405. (44) Rühling, Å.; Tyler, G. Changes in atmospheric deposition rates of heavy metals in Sweden. Water Air Soil Pollut.: Focus 2001, 1, 311–323.

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