Mercury Distribution across 14 U.S. Forests. Part II: Patterns of Methyl

Article pubs.acs.org/est

Mercury Distribution across 14 U.S. Forests. Part II: Patterns of Methyl Mercury Concentrations and Areal Mass of Total and Methyl Mercury Daniel Obrist*,† †

Division of Atmospheric Sciences, Desert Research Institute, 2215 Raggio Parkway, Reno, Nevada 89512 S Supporting Information *

ABSTRACT: This study characterized distribution patterns of monomethyl mercury (MeHg) and areal mass of total mercury (THg) and MeHg across U.S. upland forests. MeHg concentrations increased from surface litter (average: 0.14 μg kg−1) to intermediate (0.47 μg kg−1) and deeper, decomposed litter (1.43 μg kg−1). MeHg concentrations were lower in soils (0.10 μg kg−1 at 0−20 cm depth; 0.06 μg kg−1 at >20 cm depth). Ratios of MeHg to THg were higher in litter compared to soils. In soils, MeHg concentrations positively correlated with THg across all sites, and MeHg concentrations also increased with C content and latitude. THg areal mass ranged from 41.6 g ha−1 to 268.8 g ha−1. Largest THg mass at all sites was sequestered in soils (average of 91%), followed by litter (8%) and aboveground biomass (20 cm depth for comparison across all sites (note that soils were separated by depth, not by horizons, and hence different horizons may be present for respective depth intervals). All samples were collected following trace-metal sampling protocols (Appendix to EPA Method 1631: Total mercury in tissue, sludge, sediment, and soil by acid digestion 5922

dx.doi.org/10.1021/es2045579 | Environ. Sci. Technol. 2012, 46, 5921−5930

Environmental Science & Technology

Article

Figure 1. (A) Bar plots showing average MeHg concentrations observed in litter and soil layers across the 12 forest sites. Error bars represent standard errors of 12 sites. (B) Bar plots showing average ratios of MeHg/THg (in percent) for the same soil and litter layers. C. Bar plots showing average MeHg/C ratios for the same soil and litter layers. Bars labeled with different letters are statistically different (P < 0.05) based on Wilcoxon signed-rank tests; bars with identical letters are not statistically different.

to 0.75 μg kg−1 (averages across multiple species or sites). One of our sites, a Douglas Fir forest in Washington State, showed enhanced MeHg levels in Oe and Oa litter layers compared to all other sites (2.9 and 9.0 μg kg−1, respectively; Table 1); this site was also characterized by very high THg levels in these horizons (238 and 361 μg kg−1, respectively). In soils, MeHg concentrations averaged 0.10 μg kg−1 for the 0−20 cm depth layer and 0.06 μg kg−1 for below 20 cm depth, which is at the low end of levels reported by others (averages between 0.14 to 1.19 μg kg−1). Across the 12 sites, soil MeHg/THg ratios averaged 0.23% for depth of 0−20 cm and 0.22% for below 20 cm depth (Table 1), and as such were also at the low end of other observations (Table 2). For example, Grigal 7 summarized data from various watershed studies and reported soil MeHg/THg ratios up to 5%, with most observations (>70%) below 1% and a mean ratio of 0.6%. Generally, lower MeHg levels in our study may be due to sites being at temperate locations, while many previous studies were conducted in more northern forests that also were higher in THg (most THg concentrations >100 μg kg−1, as compared to our sites; Tables 1 and 2). Many of these sites were likely wetter and cooler and, as such, possibly more subject to watersaturated conditions. Not surprisingly, soil MeHg concentrations across the upland forests of our study were much lower than reported for most wetland soils which are subject to frequent in situ net methylation processes.24,25,30 MeHg concentrations differed between litter and soil layers (Figure 1; Table S1 of the SI). In litter, MeHg concentrations increased from surface litter horizons toward deeper, more decomposed layers, increasing in the following order: Oi (undecomposed surface litter) < Oe (partially decomposed litter) < Oa (strongly decomposed, humic substance) litter. Differences were statistically significant between Oi and Oe horizons (but only at the 10% level: P = 0.07) and between Oi

measured. On the basis of availability of biomass, litter, and soil inventories at the sites, different methods were used to scale up measured THg and MeHg concentrations to estimate areal mass of THg (g ha−1) and MeHg (mg ha−1; for details of calculations, see Table S3 of the SI). In general, biomass (dry mass) and aboveground C inventories assessed by respective site investigators were used to calculate biomass Hg. For litter THg and MeHg mass, we performed quantitative litter sampling at most sites using 20 cm diameter rings or used litter inventories where available. To determine soil mass, we used soil density, total soil mass, and soil C pools estimated by site investigators. Where available, areal mass was estimated using plot-level (n = 4) inventory data multiplied by plot-level Hg concentrations and Hg/C ratios; where inventories were only available at the site level (n = 1), we multiplied plot-level Hg concentrations and Hg/C ratios by site-level inventory data. We performed statistical tests and regression analyses using the statistics and data analysis program STATA (Version 9.1, Statacorps, College Station, Texas). All regressions discussed in text and figures were statistically significant (P < 0.05) unless noted otherwise. Statistical differences of MeHg concentrations, MeHg/C ratios, and MeHg/THg across litter and soil layers were evaluated using Wilcoxon signed-rank tests and considered statistically significant when P < 0.05.

■

RESULTS AND DISCUSSION MeHg Concentrations in Litter and Soils. MeHg concentrations in litter across the 12 forest sites (Figure 1; Table 1) were in a similar range to those in upland forests reported by others (Table 2). Litter MeHg concentrations across our sites averaged 0.14 μg kg−1 for surface Oi litter horizons, 0.47 μg kg−1 for intermediate Oe litter horizons, and 1.43 μg kg−1 for decomposed Oa litter horizons, while litter and litterfall MeHg concentrations at other sites ranged from 0.18 5923



Article

Higher MeHg concentrations in litter compared to underlying soils in upland sites have been reported by others as well.23,31 Branfireun et al. 31 attributed elevated MeHg concentrations in the O horizon to a possible combination of accumulation of atmospheric deposition of MeHg and in situ methylation during wet conditions. The generally higher MeHg concentrations in litter compared to soils across the 12 forest sites may be driven in fact by atmospheric inputs given that enhanced litter MeHg levels also were observed at drier sites including locations in Nevada and California (e.g., sites S4 and S6)that may not have sufficiently wet conditions to facilitate significant in situ methylation during most of the year. In contrast to this study, Schwesig et al. (Table 123,32) showed decreasing MeHg concentrations in decomposed litter in a Norway pine forest compared to surface Oi litter, and the authors proposed that reasons for this include immobilization or transformation of MeHg like demethylation in the upper forest floor. Similar depth patterns as observed for MeHg concentrations in litter was previously observed for THg concentrations and THg/C ratios, 9 although depth increases were more pronounced for THg. For example, MeHg concentrations between Oi and Oa horizons increased by a factor of 1.6, compared to a factor 5.2 for THg. As a result, MeHg/THg ratios decreased with depth from litter to soil layers in nine of the 12 sites (Table 1; Figure 1), averaging 0.39% to 0.48% in litter layers but only 0.23% (0−20 cm) and 0.22% (below 20 cm) in soils across the 12 sites. Surface litter MeHg/THg were significantly higher compared to soil MeHg/THg which could be due to atmospheric inputs accumulating in surface layers, as discussed above. For example, MeHg/THg in precipitation has been reported up to about 5% in open precipitation33−35 and between 0.6 and 2.0% in throughfall deposition,35 possibly leading to higher MeHg/THg ratios in surface layers. Decreasing MeHg/THg ratios from litter to soils also could be caused by MeHg transformation, such as demethylation in deeper litter layer or soils as proposed by Schwesig and Matzner.32 A mass balance study performed by St. Louis et al.36accounting for litterfall, throughfall inputs, watershed exports, as well as rate of MeHg accumulationsuggested that atmospheric deposition exceeded MeHg accumulation rates in soils, and about 1 mg MeHg ha−1 yr−1 may be lost, possibly via demethylation. Microbial demethylation of MeHg occurs under aerobic conditions,27,37 and thin layers of upland soils may favor presence of demethylating bacteria.38 Reduction of MeHg possibly by photochemical processesand subsequent evasion losses as elemental, inorganic Hg have been reported on foliage by Mowat et al.,39 but it is unclear if such losses also occur from litter and soils. Differences in MeHg/THg ratios in litter and soil also could be driven by different mobility and retention capacity for MeHg and THg in litter and soils; for example, Schwesig et al.23,32 attributed declining MeHg/THg ratios from surface litter to decomposed litter and into mineral soils to THg more efficiently percolating through forest litter. They showed higher retention of MeHg compared to THg in litter horizons, with only 19% of MeHg in surface deposition that reached the mineral soils through surface litter horizons, while for THg that amount was 60%. Finally, litter and soils can be sites of in situ MeHg production,7,27 particularly under anoxic conditions and in the presence of sulfate-reducing bacteria;26,40 and in situ methylation could be contributing to MeHg/THg depth patterns. Litter decomposition studies have shown, however, that mass and concentration gains of MeHg in litter

Table 1. Concentrations of THg, MeHg, and MeHg/THg Ratios (in Percent) Shown for Litter and Soil Horizons for All Sites site location (dominant species)

mercury (in μg kg−1)

litter Oi

litter Oe

litter Oa

S1

Gainesville, FL (Pine) Oak Ridge, TN (Oak, Maple, Hickory)

S3

Ashland, MO (Oak/Hickory)

S4

Little Valley, NV (Jeffrey Pine)

S6

Marysville, CA (Blue Oak)

S7

Truckee, CA (Jeffrey Pine)

S9

Niwot Ridge, CO (Fir, Spruce)

S10

Hart, MI (Sugar Maple)

S11

Bartlett, NH (Maple, Beech)

S12

Howland, ME (Spruce, Hemlock)

S13

Thompson Forest, WA (Douglas Fir)

S14

Thompson Forest, WA (Red Alder)

33 0.05 0.15 87 0.23 0.27 36 0.04 0.11 35 0.09 0.27 36 0.29 0.82 21 0.08 0.41 28 0.25 1.06 44 0.14 0.30 26 0.04 0.17 33 0.09 0.28 51 0.25 0.50 24 0.15 0.64 38 0.14 0.42 34 0.12 0.29

52 0.06 0.12 202 0.29 0.14 64 0.17 0.27 82 0.31 0.40 55 0.18 0.30 48 0.20 0.42 82 0.36 0.43 89 0.24 0.27 81 0.26 0.33 73 0.22 0.32 238 2.90 1.22 94 0.48 0.50 97 0.47 0.39 81 0.25 0.33

N/A

S2

THg MeHg % MeHg THg MeHg % MeHg THg MeHg % MeHg THg MeHg % MeHg THg MeHg % MeHg THg MeHg % MeHg THg MeHg % MeHg THg MeHg % MeHg THg MeHg % MeHg THg MeHg % MeHg THg MeHg % MeHg THg MeHg % MeHg THg MeHg % MeHg THg MeHg % MeHg

site

average, all 12 sites

median, all 12 sites

466 0.45 0.09 N/A

N/A

N/A

136 0.17 0.12 211 0.22 0.11 135 0.23 0.17 170 0.15 0.09 251 0.19 0.07 361 9.06 2.52 160 0.99 0.67 236 1.43 0.48 190 0.23 0.12

soil 0−20 cm

soil > 20 cm

10 0.08 0.60 222 0.11 0.06 42 0.04 0.10 9 0.04 0.44 46 0.03 0.07 33 0.18 0.49 19 0.03 0.18 31 0.05 0.16 44 0.06 0.16 54 0.09 0.19 167 0.25 0.15 158 0.20 0.13 70 0.10 0.23 43 0.07 0.16

9 0.10 0.88 69 0.03 0.04 40 0.01 0.02 7 0.04 0.53 27 0.01 0.05 23 0.04 0.16 26 0.02 0.09 25 0.02 0.08 39 0.09 0.38 43 0.06 0.20 137 0.09 0.07 129 0.16 0.12 48 0.06 0.22 33 0.04 0.11

and Oa horizons (P = 0.04) using data from all sites, but differences were not significant between Oe and Oa horizons. Although these depth patterns were not fully consistent across sites, 11 of the 12 sites showed Oe litter with MeHg concentrations that were higher than Oi litter concentrations. In seven of the eight sites where Oa litter horizons were present, Oa-litter MeHg levels exceeded those of Oi litter. In soils, MeHg concentrations were significantly lower compared to litter, and MeHg concentrations in the top 0−20 cm were approximately double (significantly higher) than those below 20 cm depth. When MeHg was standardized per unit of organic C (i.e., MeHg/C ratios), however, ratios increased from surface to deeper layers, with statistically significant differences observed between surface litter (Oi and Oe) and soils. 5924



Article

Table 2. THg, MeHg, and MeHg/THg Ratios (in Present) of Upland Forest Sites from Published Studies species, forest type, location Cinnamomum, Mt. Leigong, China Rhododendron, Mt. Leigong, China Fargesia, Mt. Leigong, China Pine, Ontario, Canada Old growth forest, Ontario, Canada Jack Pine, Ontrio, Canada Old growth forest, Ontario, Canada Deciduous forest, Ontraio, Canada Jack Pine (>75yr), Ontario, Canada Balsam Fir (22yr), Ontario, Canada Pine/Birch/Alnus (22yr), Ontario, Canada Pine/Birch (22yr), Ontario, Canada Pine/Birch (15yr), Ontario, Canada Pine (15yr), Ontario, Canada Norway pine, Coulissenhieb, Germany

Norway pine, Lehstenbach, Germanya

Sessile oak, European beech, Steinkreuz, Germanya

Gåardsjön, Sweden

Norway pine, Coulissenhieb, Germanya

Norway pine, Weidenbrunnen, Germanya

Scots pine-Norway Pine, Sweden

THg μg kg−1

MeHg μg kg−1

Me Hg/THg ratio (in percent)

litterfall litterfall litterfall Average litterfall litterfall litterfall litterfall litterfall Average litterfall litterfall litterfall

106 57 110 91 29.5 48.4 28.9 69.2 29.3 41 79 53 29

0.80 0.44 0.72 0.65 0.22 0.37 0.23 0.29 0.49 0.32 0.24 0.06 0.19

0.75 0.77 0.65 0.73 0.75 0.76 0.80 0.42 1.67 0.88 0.30 0.11 0 66

litterfall litterfall litterfall Average litter-Oi litter-Oe litter-Oa Average soil - 0 cm soil - 10 cm soil - 20 cm soil - 30 cm soil - 50 cm Average soil - 0 cm soil - 5 cm soil - 15 cm soil - 25 cm soil - 35 cm soil - 45 cm soil - 60 cm Average litter 0−6 cm humus - 4−10 cm bleach - 8−13 cm peat/moss - 0−10 cm Average mineral soil, Bhf/Bf - 8−45 cm mineral soil, BC/C Average upland soil - 0 cm upland soil - 5 cm upland soil - 15 cm upland soil - 25 cm upland soil - 40 cm upland soil - 55 cm Average upland soil - 0 cm upland soil - 5 cm upland soil - 15 cm upland soil - 25 cm upland soil - 40 cm upland soil - 55 cm Average site M: soil above GW table site K: soil above GW table

30 25 33 42 123 377 495 332 500 120 150 260 100 226 150 100 60 45 35 30 20 63 146 109 28 215 125 46 22 34 500 110 175 260 100 20 194 360 120 175 185 190 95 188 85 77

0.15 0.06 0.40 0.18 0.94 0.55 0.24 0.58 0.40 0.30 0.45 1.70 0.10 0.59 0.80 0.30 0.20 0.15 0.10 0.12 0.10 0.25 0.97 0.53 0.04 1.35 0.72 0.30 0.06 0.18 0.30 0.25 0.35 1.80 0.10

0.50 0.24 1.21 0.50 0.76 0.15 0.05 0.32 0.08 0.25 0.30 0.65 0.10 0 28 0−53 0.30 0.33 0.33 0.29 0.40 0.50 0.38 0.66 0.49 0.13 0.63 0.48 0.65 0.29 0.47 0.06 0.23 0.20 0.69 0.10

0.56 0.30 0.15 0.20 0.50 0.45 0.20 0.30 2.86 0.60

0.26 0.08 0.13 0.11 0.27 0.24 0.21 0.17 3.36 0.78

ecosystem component

5925

reference Fuetal., 2010

Graydon et al. 2008

St. Louis et al., 2001

Schwesigetal. 1999/Schwesigand Matzner, 2001

Schwesig and Matzner 2000

Munthe et al., 1998

Schwesigetal., 1999

Skyllberg et al., 2003



Article

Table 2. continued species, forest type, location

Upland soils, Ontario, Canada

Adirondack forest soil, U.S. a

ecosystem component site V: soil above GW table Average upland soil organic upland soil mineral Average active pool (forest floor) less active pool (mineral soil)

THg μg kg−1

MeHg μg kg−1

Me Hg/THg ratio (in percent)

15 59

0.11 1.19 0.23 0.04 0.14 94b 85b

0.73 1.63

15 546b 28 719b

reference

Branfirueun and Roulet 2002

0.60 0.30

Selvendiranetal., 2008

Values approximated from graphical representation of data. bValures are Hg/C element ratios (in μg kg−1)

only occurred when conditions were saturated,41,42 while in unflooded, upland conditionssuch as found across the 12 sites of this studyMeHg losses are generally observed during litter decomposition.42 Spatial Patterns and Correlations between MeHg and THg. Patterns of MeHg across the 12 sites showed similarities with the distribution of THg in soils and, to a lesser degree, in litter layers (Figure 2A; Table S2 of the SI). In soils, THg and

is highly complex, includes different methylation and demethylation pathways, and is dependent on microbial and environmental conditions including temperature, pH, oxygen availability, redox potential, humic acids, and many other factors.27 In terrestrial environments, the speciation of Hg is further related to hydrologic conditions,31 affected by differential mobility and retention of species in litter and soils 32 as well as influenced by atmospheric deposition patterns.35 It was surprising to find simple linear correlations between THg and MeHg across these 12 sites, particularly since the sites show profound differences in climate, soil type, geology, soil moisture regimes, and ecological properties (see Table S1 of the SI). It is possible that the positive relationships between MeHg and THg may be due to indirect effects, including a common affinity of THg and MeHg to soil C, which shows strong positive correlations to both MeHg and THg (see below). In a companion study, soil THg concentrations across these forests sites showed positive correlations to soil organic C, latitude, annual precipitation, and clay content.9 Similar although less consistentrelationships to some of these variables also were found for MeHg concentrations. Most notably, MeHg was related to soil C across the 12 forest sites (r2 of 0.50 and 0.34 for soils 0−20 and >20 cm depth, respectively; Table S2 of the SI). MeHg, similar to THg, shows strong associations with soil organic matter,44 which is likely due to a common affinity to reduced organic sulfur groups in organic matter.45 Previous studies reported correlations between MeHg, THg, and organic matter content across soil depth profiles.46 Regressions between MeHg and soil C found across the 12 sites in our study, however, are in contrast to Skyllberg et al.47 who reported no correlation between THg and MeHg in soil transects in Sweden, and also did not observe correlation between MeHg and organic C content. These differences may be attributed to the fact that the study by Skyllberg et al. included both upland sites and near stream bank locations, which likely have very different methylation potentials, leading to gradients in MeHg levels and MeHg/ THg ratios; our study, however, compared upland forest sites only. Statistically significant linear regressions of MeHg also were found with latitude in three of five litter and soil layers (litter Oe, litter Oa, and soils 0−20 cm; Table S2 of the SI). Possibly, such correlations could be due to colder and wetter climates leading to higher soil moisture levels at northern sites. Another interesting observation was that across the 12 sites, the ratio of MeHg to THg significantly correlated with soil texture (Figure 2B). MeHg/THg ratios increased with increasing sand content, and decreased with increasing silt and clay content at both depth layers. THg, however, positively correlated to clay and, to a lesser degree, silt content across these same sites,9 which may be linked to sorption of THg to

Figure 2. (A) Linear regressions between MeHg and THg in soils, separated into 0−20 cm depth (filled symbols) and below 20 cm depth (open symbols). Solid line represents the regression line for the 0−20 depth, the dashed line is for soils >20 cm depth. (B) MeHg/THg ratios (in percent) as a function of soil texture (% clay, % silt, % sand), also separated into 0−20 cm depth (filled symbols, solid line) and >20 cm depth (open symbols, dashed lines).

MeHg were significantly correlated across the 12 sites; for the 0−20 cm soil depth, a linear regression between THg and MeHg showed a positive slope and coefficient of determination (r2) of 0.20; for soils below 20 cm, a linear regression showed an r2 of 0.20 (Figure 2A). In litter layers, we observed statistically significant linear regressions between THg and MeHg in the Oe litter horizon (r2 = 45), but not in the Oi or Oa layer. In forest soils, THg seems to have some predictive value for MeHg across multiple sites, with sites higher in THg showing, on average, higher levels of MeHg. In contrast to this, THg concentration in streams and catchment runoff is not considered a good predictor of MeHg concentration as evident by a lack of correlation between THg and MeHg in streams and runoff from catchments. In lakes and ponds, however, THg has been shown to correlate with MeHg within individual geographic areas.43 The speciation between THg and MeHg 5926



Article

Table 3. Summary of Estimated Areal Mass of THg and MeHg in Aboveground Biomass (THg only), Litter, And Soil and Total Areal Mass for Thg (Aboveground Biomass Plus Litter Plus Belowground for THg) and MeHg (Litter Plus Soil)a areal mass of THg (g ha‑1)

site description site S1

Gainesville, FL

S2

Oak Ridge, TNb Ashland, MO

S3 S4 S5 S6 S7 S8 S9

Little Valley, NV Little Valley, NV Marysville, CA Truckee, CA Truckee, CA

S10 Sll

Niwot Ridge, CO Hart, Ml Bartlett, NH

S12

Howland, ME

S13

Thompson Forest, WA Thompson Forest, WA

S14

dominant species

lat.

long.

elevation

aboveground biomass

litter

below ground (0−40 cm)

total

% in biomass

% in litter

% in soil

Slash/ Longleaf Pine Oak, Maple, Hickory Oak/ Hickory Jeffrey Pine

29.74

−82.22

50

0.7 (0.0)

0.4 (0.0)

49.3 (18.0)

50.6 (17.9)

1.4

0.8

97.3

35.97

−84.28

325

2.6 (0.3)

5.5 (1.5)

395.0 (18.0)

402.6 (19.0)

0.7

1.4

98.1

38.73

−92.20

210

1.1 (0.3)

0.2 (0.0)

214.2 (19.7)

215.2 (19.7)

0.5

0.1

99.5

39.12

−119.93

2010

1.3 (0.5)

1.6 (0.2)

41.4 (0.5)

44.3 (1.1)

2.9

3.5

93.5

Manzanita (postfire) Blue Oak

39.12

−119.93

2010

0.0 (0.0)

1.5 (0.0)

40.1 (0.9)

41.6 (0.9)

0.1

3.7

96.2

39.25

−121.28

386

0.4 (0.1)

0.3 (0.0)

113.6 (14.2)

114.3 (14.3)

0.3

0.2

99.4

Jeffrey Pine Jeffrey Pine (postfire) Fir,Spruce

39.37 39.37

−120.16 −120.16

1767 1767

0.6 (0.1) 0.5 (0.2)

2.0 (0.9) 0.8 (0.2)

45.4 (4.8) 67.9 (4.0)

48.1 (5.3) 69.2 (3.8)

1.3 0.7

4.2 1.2

94.5 98.1

40.03

−105.55

3050

2.6 (0.2)

50.7 (8.9)

41.5 (5.8)

94.8 (14.3)

in

53.4

43.8

43.67 44.06

−86.15 −71.29

210 272

0.9 (0.1) 1.0 (0.0)

2.1 (0.5) 7.1 (0.6)

144.5 (15.9) 32.1 (5.8)

147.5 (15.5) 40.2 (5.5)

0.6 2.5

1.4 17.7

98.0 79.8

45.20

−68.74

60

0.8 (0.0)

69.8 (3.4)

83.9 (19.9)

154.6 (22.9)

0.5

45.2

54.3

47.38

−121.93

220

0.9 (0.2)

8.1 (0.5)

259.8 (15.8)

268.8 (16.0)

0.3

3.0

96.7

47.38

−121.93

220

0.2 (0.1)

9.0 (0.9)

235.0 (19.7)

244.2 (19.1)

0.1

3.7

96.2

1.0

11.4

1

8

91

site location

Sugar Maple Maple, Beech Spruce, Hemlock Douglas Fir, WA Red Alder, WA

average

site

site location

S1

Gainesville, FL

S2

Oak Ridge, TNb

S3 S4 S5

Ashland, MO Little Valley, NV Little Valley, NV

S6 S7 S8

Marysville, CA Truckee, CA Truckee, CA

S9

Niwot Ridge, CO Hart, Ml Bartlett, NH Howland, ME

S1O Sll S12 S13 S14

Thompson Forest, WA Thompson Forest, WA

dominant species

126.0 138.3 areal mass of MeHg (mg ha−1)

% in litter

% in soil

(187.0)

0.2

99.8

182.1

(109.4)

3.5

96.5

107.4 (5.8) 203.1 (55.0)

107.9 208.1

(5.9) (54.0)

0.5 2.4

99.5 97.6

1.1 (0.6) 3.4 (0.1)

74.9 (23.9) 165.4 (53.7)

76.0 168.8

(23.3) (53.7)

1.4 2.0

98.6 98.0

3050

77.5 (10.8)

101.8 (40.9)

179.3

(51.8)

43.2

56.8

−86.15 −71.29 −68.74

210 272 60

3.7 (0.2) 6.7 (2.1) 53.1 (8.0)

135.3 (9.7) 68.056.3) 169.8 (25.2)

139.1 74.8 222.9

(9.4) (58.4) (34.3)

2.7 9.0 23.8

97.3 91.0 76.2

47.38

−121.93

220

182.9 33.1

260.4 (20.1)

443.2

(13.0)

41.3

58.7

47.38

−121.93

220

56.4 (38.0)

245.1 (21.2)

301.5

(59.2)

18.7

81.3

33.1

171.9

205.0

litter

below ground (0−40 cm)

lat.

long.

elevation

Slash/Longleaf Pine Oak, Maple, Hickory Oak/Hickory Jeffrey Pine Manzanita (postfire) Blue Oak Jeffrey Pine Jeffrey Pine (postfire) Fir,Spruce

29.74

−82.22

50

0.5 (0.0)

356.1 (187.0)

356.6

35.97

−84.28

325

6.3 (0.9)

175.8 (108.6)

38.73 39.12 39.12

−92.20 −119.93 −119.93

210 2010 2010

0.5 (0.1) 5.0 (1.0)

39.25 39.37 39.37

−121.28 −120.16 −120.16

386 1767 1767

40.03

−105.55

Sugar Maple Maple, Beech Spruce, Hemlock Douglas Fir, WA Red Alder, WA

43.67 44.06 45.20

average

litter plus soil

12.4

87.6

a

Numbers in parentheses represent estimated standard errors in pools, based on observed concentration variability of four (and two for MeHg) sampled plots variability in observed biomass, soil, and litter inventory (where available). bPolluted site.

clay minerals.48 The correlations of MeHg/THg ratios to soil texture were unlikely codetermined by C content, latitude, longitude, precipitation, or elevation, all of which were unrelated to soil texture. Possible reasons for these patterns include that soil textures lead to differential retention of THg and MeHg or that soil texture affects other processes such as

MeHg methylation or demethylation processes, for example via effects on soil moisture. Mass of THg and MeHg Across Forest Sites. Total areal mass of THg differed by a factor of 6 across the 13 sites we considered unpolluted, ranging from 41.6 g THg ha−1 in a Jeffrey Pine forest in the Sierra Nevada mountain range to 268.8 g THg ha−1 in a Douglas fir forest in Washington State 5927



Article

characterized by particularly high sand content, showed the second highest MeHg mass while it only ranked 10th in regard to THg mass. The regression slopes suggest an increase in areal mass of Hg, on average, of 10.6 g ha−1 per degree latitude for THg and 20 μg ha−1 per degree latitude for MeHg across these sites. These results indicate that latitudinal gradients in concentrations translate into similar gradients in areal mass, and that higher mass storage of THg and MeHg can be expected in northern forests in the contiguous United States. A limitation of our estimates of areal mass is that they were standardized to the top 0−40 cm soil depth and do not account for potentially different soil depths; this may not be a major issue given that most Hg is considered to accumulate in surface reservoirs.7,28 Spatial patterns of atmospheric Hg concentrations and rates of depositionfor example, as assessed by the EPA Toxics Release Inventory for Hg,50 Hg wet deposition based on National Atmospheric Deposition Program maps,51 and atmospheric gaseous and oxidized Hg levels and depositions based on models17are not known to show latitudinal gradients and, hence, are unlikely to directly cause these south−north gradients. Climate gradients can affect mean residence times of C in soils and terrestrial ecosystems,52 and this may be linked to observed latitudinal gradients of THg and MeHg concentrations and areal mass. We previously suggested that increased THg concentrations in northern sites may be due to older C pools that have accumulated a historic legacy of atmospheric pollution.9 No corresponding latitudinal gradients of C pools across these sites were observed, but it is important to note that C pool sizes are both a function of C residence time and site productivity (e.g., litter inputs52,53). An alternative explanation is that latitudinal patterns are due to differences in vegetation types; for example, most of the northern forests were dominated by coniferous species which generally were characterized by higher litterfall Hg concentrations29,54,55 and higher throughfall deposition.56−59 A detailed comparison of two sites located in Washington State 29 showed pronounced differences in THg and MeHg levels between adjacent coniferous and deciduous forests; at these sites, however, concentration discrepancies did not translate into differences in areal mass because of counteracting differences in soil bulk densities. Another reason for latitudinal increases in THg and MeHg mass includes a potential “grasshopper” effect caused by temperature-dependent condensation and evaporation processesor possibly by different radiation environmentsthat lead to net migration of many semivolatile pollutants toward colder, higher latitudes.60,61

(Table 3; Figure 3A). The largest THg pools were estimated for the Oak Ridge site (402.6 g THg ha−1), which is likely a

Figure 3. (A) Areal mass of THg including aboveground biomass, litter, and soils of the 14 forest sites; mass was estimated based on detailed aboveground biomass inventories as well as litter and soil mass and C pool estimates. One site considered affected by local pollution is marked with an open symbol and is not included in the regression analysis. (B) Estimated areal mass of MeHg in soils and litter (no aboveground samples were analyzed for MeHg) at the 12 sites where MeHg was measured. The site affected by local pollution and a site characterized by unusually high sand content are marked with an open symbol and are not included in regression analysis.

result of nearby sources of atmospheric emissions9 and hence higher deposition and accumulation in soils relative to the other sites. In all sites, the largest THg mass was found in soils, which on average accounted for 91% (>90% in 11 of the 14 sites) of the total areal mass. This is related to much higher bulk densities of soils compared to litter layers. On average, litter horizons accounted for 8% of total mass, although higher litter Hg proportions were found in two forest sites with extensive and deep litter layers [a northeastern coniferous site in Maine (45% of Hg found in litter) and a high altitude Rocky Mountain site (53% of Hg found in litter)]. Further, aboveground standing biomass accounted, on average, for less than 1% of the areal mass, with highest proportions in a dense, Rocky Mountain coniferous forest (2.7%). Although significant C stores are associated in aboveground biomass in forests, the small mass of THg therein is due to the very low concentrations of THg found in bole wood. These patterns show that soils are by far the most important compartment for THg storage in these forests. Combined litter plus soil MeHg mass ranged from 75 to 443 mg ha−1 across the 12 investigated sites (Table 3; Figure 3B). Similar to THg, we found the largest mass in soils which on average accounted for 88% of areal mass and exceeded 90% in 8 of the 12 sites, with litter accounting for the rest. Schwesig and Matzner 49 reported that the proportion of MeHg in litter accounted for 5−13% of the soil storage (0−60 cm depth), and Munthe et al. 44 showed that litter accounted for 11−30% of MeHg storage in Swedish soils. For both THg and MeHg, areal mass showed latitudinal trends (Figure 3): a linear regression of THg mass to latitude showed an r2 of 37% (without the polluted Tennessee site S2; Figure 3A), and of 39% for MeHg (without the polluted site and sandy Florida site S1; Figure 3B). The Florida site,

■

ASSOCIATED CONTENT

S Supporting Information *

Detailed description of sites (in text, Figure S1 and Table S1), results of linear regression analyses of MeHg concentrations to soil C, latitude, and precipitation (Table S2), and scaling protocols used to scale up areal mass of THg and MeHg at each site. This material is available free of charge via the Internet at http://pubs.acs.org.

■

AUTHOR INFORMATION

Corresponding Author

*Phone: (775) 674-7008; e-mail: [email protected]. Notes

The authors declare no competing financial interest. 5928



■

Article

Louis, V. L. S.; Tate, M. T. Whole-ecosystem study shows rapid fishmercury response to changes in mercury deposition. Proc. Natl. Acad. Sci. U.S.A. 2007, 104 (42), 16586−16591. (14) Pokharel, A. K.; Obrist, D. Fate of mercury in tree litter during decomposition. Biogeosciences 2011, 8 (9), 2507−2521. (15) Obrist, D.; Faïn, X.; Berger, C. Gaseous elemental mercury emissions and CO2 respiration rates in terrestrial soils under controlled aerobic and anaerobic conditions. Sci. Total Environ. 2010, 408, 1691− 1700. (16) Obrist, D. Atmospheric mercury pollution due to losses of terrestrial carbon pools? Biogeochemistry 2007, 85 (2), 119−123, DOI: 10.1007/s10533-007-9108-0. (17) Selin, N. E.; Jacob, D. J.; Yantosca, R. M.; Strode, S.; Jaegle, L.; Sunderland, E. M. Global 3-D land-ocean-atmosphere model for mercury: Present-day versus pre-industrial cycles and anthropogenic enrichment factors for deposition. Global Biogeochem. Cycles 2008, 22, 2. (18) 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 (5071), 784−787. (19) Driscoll, C. T.; Han, Y. J.; Chen, C. Y.; Evers, D. C.; Lambert, K. F.; Holsen, T. M.; Kamman, N. C.; Munson, R. K. Mercury contamination in forest and freshwater ecosystems in the Northeastern United States. Bioscience 2007, 57 (1), 17−28. (20) Selvendiran, P.; Driscoll, C. T.; Bushey, J. T.; Montesdeoca, M. R. Wetland influence on mercury fate and transport in a temperate forested watershed. Environ. Pollut. 2008, 154 (1), 46−55. (21) Selvendiran, P.; Driscoll, C. T.; Montesdeoca, M. R.; Bushey, J. T., Inputs, storage, and transport of total and methyl mercury in two temperate forest wetlands. J. Geophys. Res.-Biogeosci. 2008, 113. (22) Branfireun, B. A.; Roulet, N. T.; Kelly, C. A.; Rudd, J. W. M. In situ sulphate stimulation of mercury methylation in a boreal peatland: Toward a link between acid rain and methylmercury contamination in remote environments. Global Biogeochem. Cycles 1999, 13 (3), 743− 750. (23) Schwesig, D.; Ilgen, G.; Matzner, E. Mercury and methylmercury in upland and wetland acid forest soils of a watershed in NE-Bavaria, Germany. Water, Air, Soil Pollut. 1999, 113 (1−4), 141−154. (24) St. Louis, V. L.; Rudd, J. W. M.; Kelly, C. A.; Beaty, K. G.; Bloom, N. S.; Flett, R. J. Importance of wetlands as sources of methylmercury to boreal forest ecosystems. Can. J. Fish. Aquat. Sci. 1994, 51 (5), 1065−1076. (25) Hurley, J. P.; Benoit, J. M.; Babiarz, C. L.; Shafer, M. M.; Andren, A. W.; Sullivan, J. R.; Hammond, R.; Webb, D. A. Influences of watershed characteristics on mercury levels in Wisconsin rivers. Environ. Sci. Technol. 1995, 29 (7), 1867−1875. (26) Verta, M.; Matilainen, T.; Porvari, P.; Niemi, M.; Uusl-Rauva, A.; Bloom, N. S., Methylmercury sources in boreal lake ecosystems. In Mercury Pollution: Integration and Synthesis; Watras, C. J., Huckabee, J. W., Eds.; CRC Press: Monterey, CA, 1994; pp 119−136. (27) Ullrich, S.; Tanton, T.; Abdrashitova, S. Mercury in the aquatic environment: A review of factors affecting methylation. Crit. Rev. Environ. Sci. Technol. 2001, 31 (3), 241−293. (28) Obrist, D.; Johnson, D. W.; Lindberg, S. E. Mercury concentrations and pools in four Sierra Nevada forest sites, and relationships to organic carbon and nitrogen. Biogeosciences 2009, 6 (5), 765−777. (29) Obrist, D.; Johnsen, D. W.; Edmonds, R. L., Effects of vegetation type on mercury concentrations and pools in two adjacent coniferous and deciduous forests. J. Plan. Nutr. Soil Sci., in press. (30) Heyes, A.; Moore, T. R.; Rudd, J. W. M.; Dugoua, J. J. Methyl mercury in pristine and impounded boreal peatlands, experimental Lakes Area, Ontario. Can. J. Fish. Aquat. Sci. 2000, 57 (11), 2211− 2222. (31) Branfireun, B. A.; Roulet, N. T. Controls on the fate and transport of methylmercury in a boreal headwater catchment, northwestern Ontario, Canada. Hydrol. Earth Syst. Sci. 2002, 6 (4), 783−794.

ACKNOWLEDGMENTS I’d like to thank R. Bracho, J.J. Battles, D. B. Dail, R. L. Edmonds, B. Evans, B. Gynea, K. Hosman, P. Micks, R.K. Monson, P. Mulhoollan, S.V. Ollinger, S.G. Pallardy, M. Prater, K.S. Pregitzer, A. Richardson, D.E. Todd, and R. Wenk for facilitating site access and providing biomass and soil inventories of the research sites; C. Berger, J. Dagget, R. Higgins, S. Lee, G. Marty, A. Pierce, A. Pokharel, and S. Vadwalas for laboratory analyses and field sampling; L. Arnone for help with data analyses and figure and table editing; D.W. Johnson, Y. Luo, and S.E. Lindberg who served as Co-Principal Investigators of the study and provided input; and C. Moore and R. Kreidberg for editorial comments on an earlier draft of this manuscript. This study was funded by the U.S Environmental Protection Agency (U.S. EPA) through a Science-ToAchieve-Results grant (No. R833378) and received support from the Desert Research Institute (DRI) through an Internal Project Assignment grant.

■

REFERENCES

(1) Lag, J.; Steinnes, E. Regional distribution of mercury in humus layers of Norwegian forest soils. Acta Agric. Scand. 1978, 28 (4), 393− 396. (2) Hakanson, L.; Nilsson, A.; Andersson, T. Mercury in the Swedish mor layerLinkages to mercury deposition and sources of emission. Water, Air, Soil Pollut. 1990, 50 (3−4), 311−329. (3) Faïn, X.; Obrist, D.; Pierce, A.; Barth, C.; Gustin, M. S.; Boyle, D. P. Whole-watershed mercury balance at Sagehen Creek, Sierra Nevada, CA. Geochim. Cosmochim. Acta 2011, 75, 2379−2392. (4) Hintelmann, H.; Harris, R.; Heyes, A.; Hurley, J.; Kelly, C.; Krabbenhoft, D.; Lindberg, S.; Rudd, J.; Scott, K., St; Louis, V. Reactivity and mobility of new and old mercury deposition in a boreal forest ecosystem during the first year of the METAALICUS study. Environ. Sci. Technol. 2002, 36 (23), 5034−5040. (5) Meili, M. The coupling of mercury and organic matter in the biogeochemical cycleTowards a mechanistic model for the boreal forest zone. Water, Air, Soil Pollut. 1991, 56, 333−347. (6) 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 (13), 4174−4180. (7) Grigal, D. F. Mercury sequestration in forests and peatlands: A review. J. Environ. Qual. 2003, 32 (2), 393−405. (8) Hararuk, O.; Obrist, D.; Luo, Y. Modeling the sensitivity of soil mercury storage to climate-induced changes in soil carbon pools. Biogeosciences 2012, in review. (9) Obrist, D.; Johnson, D.; Lindberg, S.; Luo, Y.; Hararuk, O.; Bracho, R.; Battles, J.; Dail, D.; Edmonds, R.; Monson, R.; Ollinger, S.; Pallardy, S.; Pregitzer, K.; Todd, D. Mercury distribution across 14 U.S. forests. Part I: Spatial patterns of concentrations in biomass, litter, and soils. Environ. Sci. Technol. 2011, 45, 3974−3981. (10) Ericksen, J. A.; Gustin, M. S.; Xin, M.; Weisberg, P. J.; Fernandez, G. C. J. Air−soil exchange of mercury from background soils in the United States. Sci. Total Environ. 2006, 366, 851−863. (11) Fritsche, J.; Obrist, D.; Zeeman, M. J.; Conen, F.; Eugster, W.; Alewell, C. Elemental mercury fluxes over a sub-alpine grassland determined with two micrometeorological methods. Atmos. Environ. 2008, 42 (13), 2922−2933. (12) Ericksen, J. A.; Gustin, M. S.; Lindberg, S. E.; Olund, S. D.; Krabbenhoft, D. P. Assessing the potential for re-emission of mercury deposited in precipitation from arid soils using a stable isotope. Environ. Sci. Technol. 2005, 39 (20), 8001−8007. (13) Harris, R. C.; Rudd, J. W. M.; Almyot, M.; Babiarz, C. L.; Beaty, K. G.; Blanchfield, P. J.; Bodaly, R. A.; Branfireun, B. A.; Gilmour, C. C.; Graydon, J. A.; Heyes, A.; Hintelmann, H.; Hurley, J. P.; Kelly, C. A.; Krabbenhoft, D. P.; Lindberg, S. E.; Mason, R. P.; Paterson, M. J.; Podemski, C. L.; Robinson, A.; Sandilands, K. A.; Southworth, G. R.; 5929



Article

trends and loss mechanisms. Environ. Sci. Technol. 2010, 44 (15), 5901−5907. (51) NADP National Atmospheric Deposition Network (NADP) Mercury Deposition Network. http://nadp.sws.uiuc.edu/MDN/ (December 2010). (52) Froberg, M.; Tipping, E.; Stendahl, J.; Clarke, N.; Bryant, C. Mean residence time of O horizon carbon along a climatic gradient in Scandinavia estimated by 14C measurements of archived soils. Biogeochemistry 2011, 104 (1−3), 227−236. (53) Callesen, I.; Liski, J.; Raulund-Rasmussen, K.; Olsson, M. T.; Tau-Strand, L.; Vesterdal, L.; Westman, C. J. Soil carbon stores in Nordic well-drained forest soilsRelationships with climate and texture class. Global Change Biol. 2003, 9 (3), 358−370. (54) Hall, B.; Louis, V. Methylmercury and total mercury in plant litter decomposing in upland forests and flooded landscapes. Environ. Sci. Technol. 2004, 38 (19), 5010−5021. (55) Rasmussen, P.; Mierle, G.; Nriagu, J. The analysis of vegetation for total mercury. Water, Air, Soil Pollut. 1991, 56, 379−390. (56) Witt, E. L.; Kolka, R. K.; Nater, E. A.; Wickman, T. R. Influence of the forest canopy on total and methyl mercury deposition in the boreal forest. Water, Air, Soil Pollut. 2009, 199 (1−4), 3−11. (57) Demers, J. D.; Driscoll, C. T.; Fahey, T. J.; Yavitt, J. B. Mercury cycling in litter and soil in different forest types in the Adirondack region, New York, USA. Ecol. Appl. 2007, 17 (5), 1341−1351. (58) Poulain, A. J.; Roy, V.; Amyot, M. Influence of temperate mixed and deciduous tree covers on Hg concentrations and photoredox transformations in snow. Geochim. Cosmochim. Acta 2007, 71 (10), 2448−2462. (59) Nelson, S. J.; Johnson, K. B.; Kahl, J. S.; Haines, T. A.; Fernandez, I. J. Mass balances of mercury and nitrogen in burned and unburned forested watersheds at Acadia National Park, Maine, USA. Environ. Monit. Assess. 2007, 126 (1−3), 69−80. (60) Wania, F.; Mackay, D. Tracking the distribution of persistent organic pollutants. Environ. Sci. Technol. 1996, 30 (9), A390−A396. (61) Simonich, S. L.; Hites, R. A. Global distribution of persistent organochlorine compounds. Science 1995, 269 (5232), 1851−1854.

(32) Schwesig, D.; Matzner, E. Dynamics of mercury and methylmercury in forest floor and runoff of a forested watershed in Central Europe. Biogeochemistry 2001, 53 (2), 181−200. (33) Conaway, C. H.; Black, F. J.; Weiss-Penzias, P.; Gault-Ringold, M.; Flegal, A. R. Mercury speciation in Pacific coastal rainwater, Monterey Bay, California. Atmos. Environ. 2010, 44 (14), 1788−1797. (34) Bloom, N. S.; Watras, C. J. Observations of methylmercury in precipitation. Sci. Total Environ. 1989, 87−8, 199−207. (35) Graydon, J. A.; Louis, V. L. S.; Hintelmann, H.; Lindberg, S. E.; Sandilands, K. A.; Rudd, J. W. M.; Kelly, C. A.; Hall, B. D.; Mowat, L. D. Wet, Long-Term, and Dry Deposition of Total and Methyl Mercury in the Remote Boreal Ecoregion of Canada. Environ. Sci. Technol. 2008, 42 (22), 8345−8351. (36) St. Louis, V.; Rudd, J.; Kelly, C.; Hall, B.; Rolfhus, K.; Scott, K.; Lindberg, S.; Dong, W. Importance of the forest canopy to fluxes of methyl mercury and total mercury to boreal ecosystems. Environ. Sci. Technol. 2001, 35 (15), 3089−3098. (37) Winfrey, M. R.; Rudd, J. W. M. Environmental factors affecting the formation of methylmercury in low pH lakes. Environ. Toxicol. Chem. 1990, 9 (7), 853−869. (38) St. Louis, V. L.; Rudd, J. W. M.; Kelly, C. A.; Beaty, K. G.; Flett, R. J.; Roulet, N. T. Production and loss of methylmercury and loss of total mercury from boreal forest catchments containing different types of wetlands. Environ. Sci. Technol. 1996, 30 (9), 2719−2729. (39) Mowat, L. D.; St. Louis, V. L.; Graydon, J. A.; Lehnherr, I. Influence of forest canopies on the deposition of methylmercury to boreal ecosystem watersheds. Environ. Sci. Technol. 2011, 45 (12), 5178−5185. (40) Morel, F. M. M.; Kraepiel, A. M. L.; Amyot, M. The chemical cycle and bioaccumulation of mercury. Annu. Rev. Ecol. Syst. 1998, 29, 543−566. (41) Heyes, A.; Moore, T.; Rudd, J. Mercury and methylmercury in decomposing vegetation of a pristine and impounded wetland. J. Environ. Qual. 1998, 27 (3), 591−599. (42) Hall, B. D.; Louis, V. L. S. Methylmercury and total mercury in plant litter decomposing in upland forests and flooded landscapes. Environ. Sci. Technol. 2004, 38 (19), 5010−5021. (43) Kelly, C. A.; Rudd, J. W. M.; Louis, V. L.; Heyes, A. Is total mercury concentration a good predictor of methyl mercury concentration in aquatic systems? Water, Air, Soil Pollut. 1995, 80 (1−4), 715−724. (44) Munthe, J.; Lee, Y. H.;Hultberg, H.; Iverfeldt, K.; Borg, G. C.; Andersson, B. I.Cycling of mercury and methyl mercury in the Gårdsjön catchments. In Experimental Reversal of Acid Rain Effects: The Gårdsjön Roof Project; Hultberg, H.; Skeffington, R. A., Eds.; John Wiley & Sons: New York, 1998; pp 261−276. (45) Qian, J.; Skyllberg, U.; Frech, W.; Bleam, W. F.; Bloom, P. R.; Petit, P. E. Bonding of methyl mercury to reduced sulfur groups in soil and stream organic matter as determined by X-ray absorption spectroscopy and binding affinity studies. Geochim. Cosmochim. Acta 2002, 66 (22), 3873−3885. (46) Hultberg, H.; Iverfeldt, A.; Lee, Y. H., Methylmercury input/ output and accumulation in forested catchments and critical loads for lakes in Southwestern Sweden. In Mercury Pollution: Integration and Synthesis; Watras, C. J., Huckabee, J. W., Eds.; CRD Press, Inc.: Monterey, CA, 1994; pp 313−322. (47) 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 (1), 53−76. (48) Gabriel, M.; Williamson, D. Principal biogeochemical factors affecting the speciation and transport of mercury through the terrestrial environment. Environ. Geochem. Health 2004, 26 (4), 421−434. (49) Schwesig, D.; Matzner, E. Pools and fluxes of mercury and methylmercury in two forested catchments in Germany. Sci. Total Environ. 2000, 260 (1−3), 213−223. (50) Asaf, D.; Tas, E.; Pedersen, D.; Peleg, M.; Luria, M. Long-term measurements of NO3 radical at a semi-arid urban site: 2. Seasonal 5930


Mercury Distribution across 14 U.S. Forests. Part II: Patterns of Methyl

Recommend Documents