Vertical Profile Measurements of Soil Air Suggest Immobilization of

Article pubs.acs.org/est

Vertical Profile Measurements of Soil Air Suggest Immobilization of Gaseous Elemental Mercury in Mineral Soil Daniel Obrist,*,† Ashok K. Pokharel,† and Christopher Moore† †

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

ABSTRACT: Evasion of gaseous elemental Hg (Hg0g) from soil surfaces is an important source of atmospheric Hg, but the volatility and solid−gas phase partitioning of Hg0 within soils is poorly understood. We developed a novel system to continuously measure Hg0g concentrations in soil pores at multiple depths and locations, and present a total of 297 days of measurements spanning 14 months in two forests in the Sierra Nevada mountains, California, U.S. Temporal patterns showed consistent pore Hg0g concentrations below levels measured in the atmosphere (termed Hg0g immobilization), ranging from 66 to 94% below atmospheric concentrations throughout multiple seasons. The lowest pore Hg0g concentrations were observed in the deepest soil layers (40 cm), but significant immobilization was already present in the top 7 cm. In the absence of sinks or sources, pore Hg0g levels would be in equilibrium with atmospheric concentrations due to the porous nature of the soil matrix and gas diffusion. Therefore, we explain decreases in pore Hg0g in mineral soils below atmospheric concentrationsor below levels found in upper soils as observed in previous studies with the presence of an Hg0g sink in mineral soils possibly related to Hg0g oxidation or other processes such as sorption or dissolution in soil water. Surface chamber measurements showing daytime Hg0g emissions and nighttime Hg0g deposition indicate that near-surface layers likely dominate net atmospheric Hg0g exchange resulting in typical diurnal cycles due to photochemcial reduction at the surface and possibly Hg0g evasion from litter layers. In contrast, mineral soils seem to be decoupled from this surface exchange, showing consistent Hg0g uptake and downward redistributionalthough our calculations indicate these fluxes to be minor compared to other mass fluxes. A major implication is that once Hg is incorporated into mineral soils, it may be unlikely subjected to renewed Hg0g re-emission from undisturbed, background soils emphasizing the important role of soils in sequestering past and current Hg pollution loads.

■

legacy Hg0g re-emission (i.e., re-emission of previously deposited pollutants) accounts for 200 Mg a−1 from terrestrial surfaces, an additional 800 Mg a−1 from deeper soil reservoirs, and 300 Mg a−1 from snow-covered surfaces.22−25 This combined value of atmospheric Hg0g re-emission is more than two-thirds of the estimated worldwide primary (i.e., anthropogenic) Hg0g emissions of 1900 Mg a−1.26 One particular concern is that such legacy emissions lead to continued cycling of past pollution and contribute to exposure that is delayed in time and displaced geographically from primary emission sources.27 Over soils, Hg0g exhibits a highly dynamic, bidirectional exchange behavior (i.e., emission and deposition), and soils can serve both as net sources and net sinks for atmospheric Hg0g.21,28−35 The net exchange of Hg0g over soils is controlled by a variety of factors. For example, increasing radiation and temperature generally increase surface Hg0g emissions.36−44 As

INTRODUCTION Global atmospheric distribution and deposition is the main pathway for Hg input to many remote ecosystems.1 Upland soils serve as important links between atmospheric deposition of Hg and mobilization to wetlands, streams, and lakes,2−4 where methylation and bioaccumulation5 result in exposure of wildlife and humans to this toxic pollutant.6 Soils account for more than 90% of Hg stored in terrestrial ecosystems,7−9 with top soil Hg pools (top 40 cm) estimated at 15 230 Mg in the U.S.; or, if extrapolated globally, possibly accounting for more than 300 000 Mg.10 This large pool size stems in part from natural, geologic sources, but also is due to a legacy of past anthropogenic pollution released across centuries that has accumulated in soils.11−14 Atmospheric Hg deposition is efficiently retained in upper soils, bound in particular in humus-rich layers.7,8,15 The dominant soil Hg form is divalent HgII strongly bound to organic matter15−18 through complexation by one or two thiol groups.16,19,20 Terrestrial Hg, however, also partially volatilizes into the atmosphere as gaseous elemental Hg (Hg0g) [reviews in refs 17 and 21]. Although global estimates of terrestrial sources to the atmosphere are uncertain, model simulations indicate that © 2014 American Chemical Society

Received: Revised: Accepted: Published: 2242

October 29, 2013 December 11, 2013 January 15, 2014 January 15, 2014 dx.doi.org/10.1021/es4048297 | Environ. Sci. Technol. 2014, 48, 2242−2252

Environmental Science & Technology

Article

Table 1. Soil Elemental Content (Nitrogen, Carbon, and Mercury) and Soil Texture at Measurements Sites location Blodgett

Sagehen

soil depth 7 cm 20 cm 40 cm 7 cm 20 cm 40 cm

soil element content nitrogen and carbon (in %); mercury (in μg Kg−1)

soil texture (in %)

nitrogen

carbon

C/N ratio

mercury

sand

silt

clay

0.08 0.05 0.06 0.15 0.07 0.03

2.8 1.5 1.4 3.3 1.3 1.0

37 30 26 24 17 46

43 34 32 35 25 22

58 55 55 41 38 35

25 25 25 32 34 36

18 20 20 28 29 29

soil moisture increases, soil Hg0g emission fluxes also increasein particular after moisture pulsesalthough this flux response may be different in prewetted soils and under high moisture regimes.45−50 Increasing substrate Hg concentrations strongly increase Hg0g emissions,17,41,51−53 and soil oxygen availability, soil redox potential,54,55 as well as the presence of organic and humic matter affect soil Hg0g fluxes.56,57 Biological processes also may be linked to soil Hg0g emission fluxes.58,59 Further, air concentrations play a role in soil Hg0g exchange with laboratory studies showing the presence of compensation points, defined as an air Hg0g concentration threshold where Hg0g surface fluxes change from deposition (when exposed to high air concentrations) to emissions (when exposed to low concentrations).57,60,61 Finally, evidence shows that near-surface atmospheric chemistry affects Hg0g exchange from soils as well (e.g., ozone62,63). Most of our understanding of Hg0g exchange from soils stems from direct observations at soil surfaces (i.e., by measurement of net Hg0g exchange to the atmosphere) or is based on controlled laboratory studies. Very few studies have attempted to directly characterize concentrations and the dynamics of Hg0g within intact soils in the field. To this point, we know of only two studies that measured Hg0g concentrations in uncontaminated background soils; these conducted manual extraction of pore air from soils with limited temporal and spatial measurement resolution.55,64 A few additional studies measured pore Hg0g concentrations in contaminated soils.65−67 As a result, the dynamics of subsurface soil Hg0g concentrations, the potential diffusive redistribution of Hg0g within soils, and the contributions of subsoils to surface emissions are not well understood. The goals of this study were as follows: (1) develop and test a system that allowed for time-extended measurement of pore Hg0g concentrations at multiple depths and locations in soils with sufficient sensitivity to measure concentrations in background soils; (2) measure vertical soil Hg0g profiles in background soils including temporal and spatial evolution of volatile Hg0g and responses to environmental variables; and (3) explore implications of observed pore Hg0g patterns for soil Hg cycling and soil-atmosphere exchange processes.

permeable material allowed for transport of pore air (but not of water or other solutes) from surrounding soils into the tubing. Tubes were inserted into soils months prior the onset of measurements at depths of 7, 20, and 40 cm, with two full replicate depth profiles in two different forests (Sagehen Creek forest and Blodgett forest, see below). A multiport sampler that allowed alternate sampling between different inlet lines every 45-min in soils (and then measuring the atmospheric port for four consecutive 45-min periods to reduce extraction of air volume from soils) was connected to the soil tubes and to a gaseous mercury analyzer (Model 2537; Tekran Inc., Toronto, Canada) and an infrared gas analyzer (Model 7000; LI-COR Inc. Lincoln, Nebraska, U.S.) for measurement of Hg0g and CO2 concentrations. The Tekran 2537 analyzer was modified to facilitate low flow rates to reduce disturbance effects and pore air advection using an external low-volume pump and mass flow controller allowing the reduction of flows to 50 mL min−1. Details of the system, field installation, and field sites can be found in the SI. Continuous measurement of pore Hg0g and CO2 concentrations were conducted in two pine forests in the Sierra Nevada mountains in California, U.S., at one site (Blodgett Forest) from April to May 2009 and September 2009 to January 2010 and at a second site (Sagehen) between May and November 2010. A total of 297 days of pore Hg 0 g concentration measurements were performed spanning 14 months across these two sites. Auxiliary measurements included measurements of temperature and water content at all respective soil depths (SI) and physical and chemical characterization of soils as shown in Table 1.

■

RESULTS AND DISCUSSION System Performance and Measurement of Soil Pore Trace Gas Concentrations. Concentrations of CO2 were characterized concurrently with Hg0g since CO2 soil patterns have been well documented and thereby can be used to evaluate system performance. In addition, CO2 concentration could be measured at higher temporal resolution (e.g., 1-min) compared to the low time resolution of Hg0g measurements (45 min in our system) providing additional temporal information. Examples of 1 min resolution measurements of pore CO2 concentrations are shown in Figure 1A for the Blodgett forest across a sampling sequence that includes two profiles with three soil depths each, plus a port that was used to sample atmospheric air. The figure shows distinct concentration differences when switching between sampling ports, with the time needed to equilibrate concentrations from one sampling port to the next generally between 2 and 3 min. This time delay for concentration stabilization was due to carry-over of air in the tubing and valves between the multiport unit and analyzers. These effects, however, were short compared to the 45-min sampling period at each depth and therefore had little effect on

■

METHODS Trace Gas Monitoring System to Measure Pore Hg0g Concentrations and Measurement Locations. We developed a continuous trace-gas gradient system (Supporting Information, SI Figure 1) to sequentially measure low volumes of pore air extracted from soils via gas-permeable, hydrophobic membranes placed at various depths and locations (using up to eight different sampling ports). Soil pore air was extracted using gas-permeable, hydrophobic polytetrafluoroethylene (PTFE) membrane tubing (20 cm in length; Model 032-03; International Polymer Engineering, Tempe, Arizona, U.S.). This gas2243

dx.doi.org/10.1021/es4048297 | Environ. Sci. Technol. 2014, 48, 2242−2252


Article

Figure 1. Examples of measurement sequences of Hg0g and CO2 concentrations in soil pores using the gradient system at the Blodgett forest site. A. Examples of 1 min sampling data of CO2 concentrations with high-resolution data showing fast and effective transition between sampling depths. B. Individual 45-min pore Hg0g measurements during a 30-day measurement period. Each dot is an individual measurement, with the sequence of measurements switching between different depth at replicated depth profiles at 7 cm depth (three profiles), 20 cm depth (two profiles), and 40 cm depth (two profiles), and the atmosphere. C. Corresponding figure for CO2 concentrations (45 min averages) for the same sampling period.

observed concentrations (∼5%). Most importantly, CO2 concentrations remained relatively stable during the sampling at each port, providing evidence that soil air extractions did not induce significant disturbance effects, such as advective or diffusive transport from other soil depths. This lack of disturbance also applied to sampling of the uppermost soil depth (7 cm), indicating that advection from the atmosphere also was not a problem in the uppermost layer. Observed CO2

patterns showed, as expected, strong concentration increases with soil depth as seen in Figure 1A,C (average during 45-min sampling intervals). Across all measurements, CO2 concentrations were enhanced by 455% and 399% at sites 1 and 2, respectively, from the top 7 cm to the lowest 40 cm depth (Figure 2 and SI Figure 2). Such strong CO2 diffusion gradients are typical in soils because of in-soil CO2 production by microbial carbon mineralization and roots combined with 2244



Article

Figure 2. Contour plots of soil pore Hg0g and CO2 concentrations. A. Figures showing pore Hg0g concentrations (in ng m−3) at UCB’s Sagehen Creek Field Station and Blodgett Forest Research Station. Plots were created based on daily averaged measurements combining two replicate depth profiles (7 cm, 20 cm, and 40 cm depth; and three profiles for top 7 cm depth at Blodgett forest). Atmospheric measurements were taken approximately 1 m above the surface. For contour plots, atmospheric Hg0g measurements were treated as boundary conditions for the surface (i.e., for 0 cm soil depth). B. Contour plots for CO2 concentrations for the Sagehen Creek forest and Blodgett forest and respective soil and atmospheric observations generated from the same conditions. Scale shows the natural logarithm of measured concentrations in ppm.

levels observed in the atmosphere, and that pore Hg0g levels generally decreased with depth. Across all measurements, pore Hg0g in the top 7 cm of the soil decreased to 66% and 69%, of ambient air levels, respectively, at the two sites. At a soil depth of 20 cm, the Hg0g decrease was 81% and 92% at the two sites, and 94% at both sites at 40 cm depth, compared to atmospheric levels. Pore Hg0g at these depths was significantly lower compared to atmospheric levels (based on paired t tests of daily average values; SI Figure 2). On only 4 of 297 days of measurements were concentrations at 7 cm not below levels observed in the atmosphere. At both sites, concentrations were statistically different between all soil depths as well, with concentrations decreasing with depth. The pattern of decreasing Hg0g concentrations with soil depth was relatively consistent in time and across seasons. At Blodgett forest, 79% of Hg0g concentrations were lower at the 20 cm depth compared to 7 cm, and 97% of concentrations were lower at the 40 cm depth compared to 20 cm. At Sagehen Creek forest, 81% of the concentrations were higher at 7 cm compared to 20 cm, and 68% were higher at 20 cm compared to 40 cm [note, however, that very low concentrations close to and often below detection limits were observed at this site at both 20 and 40 cm (averaging 0.09 and 0.07 ng m−3, respectively)]. In the absence of any active sink or source processes, soil Hg0g levels would be in equilibrium with ambient air

consecutive diffusive mixing with the atmosphere [see detailed discussion below, and in ref 68]. Repeated measurements both of Hg0g and CO2 concentrations (Figure 1B,C) showed that consecutive measurements at respective depths and locations were replicable, both temporally and spatially. Finally, sampling of atmospheric concentrations showed averages of 526 ± 128 ppm and 446 ± 94 ppm (for CO2) and 1.16 ± 0.35 ng m−3 and 1.64 ± 0.24 ng m−3 (for Hg0g), respectively, during the entire sampling period at the two sites (Figures 2 and SI Figure 2). For CO2, these concentrations were well within levels generally observed in forests under tree canopies.69 For Hg0g, these levels were within the range of background ambient air Hg0g concentrations.70 Somewhat lower ambient air Hg0g concentrations at Blodgett forest compared to atmospheric concentrations generally observed in the area may be attributed to flow meter calibration uncertainties (i.e., overestimating real sampling volume) or instability of flows by deploying an external flow control. Therefore, soil Hg0g concentrations at this second site may show a low bias, which we estimate at a maximum of 20%. Soil Pore Hg0g Depth Patterns. Contour and box plots of pore Hg0g and CO2 concentrations measured in mineral soil at depths of 7, 20, and 40 cm, and in the atmosphere at the two sites are shown in Figure 2 and SI Figure 2. The contour plots show that pore Hg0g concentrations were consistently below 2245



Article

below). Among clay minerals, Illite shows the highest sorption capacity and kaolinite the lowest, and the highest sorption occurs in humic acid and organic peat, followed by top soils high in organic matter. Sorption experiments show that adsorbed Hg0 is either very tightly bound to the soil matrix or transformed into nonvolatile, oxidized forms. For example, Hg0 was not lost under ambient temperatures, after placement in vacuum, and was not subject to extraction by water, KCl, CaCl2, or cation exchange.49,74,75 Another process for Hg0g immobilization is oxidation of Hg0 in soils, and Landa75 quantified that small fractions of sorbed Hg0 in soils (between 2 and 27%) were converted into HgII. Similarly, do Valle et al.77 found that tropical soils exposed to Hg0g showed significant (up to 68%) conversion of sorbed Hg0 to HgII and monovalent Hg (HgI). Schlüter et al.17 proposed that due to strong binding of HgII to organic matter, oxidation of Hg0 to HgII may proceed at redox potentials lower than its half-reaction suggests, and that Hg0g evaporating from deeper soil layers may be reoxidized in upper soil layers. Oxidation of Hg0 to HgII has been documented in particular in natural waters78−82 where oxidation is strongly photochemically driven,79,83−86 but also occurs in the dark.78,86 A third process for Hg0g immobilization may be dissolution of Hg0 in soil solution; the solubility of Hg0 in water at 20 °C is about 50 μg L−1.87 Under wet conditions and typical bulk densities observed in our soils, we calculate a potential solubility of Hg0 in soil solution in the range of 6 mg m−2, a substantial amount, potentially causing significant pore Hg0g immobilization. We assessed relationships of pore Hg0g concentrations to environmental variables in order to assess the involvement of these processes. First, pore Hg0g concentrations had no discernible diurnal patterns, which was not surprising given very limited light penetration into soils (estimated to about 2 mm88) indicating no role of photochemical processes in mineral soils. We observed no significant seasonal patterns or relationships with soil temperature either (Figure 2). Individual strong precipitation events at both sites led to increases in pore Hg0g concentrations (Figure 3), which may indicate a link between pore Hg0g immobilization and soil redox conditions. Moore and Castro55 showed that pore Hg0g levels in background soils were inversely correlated with soil redox potential, and Allen and Arsenie89 reported that in aqueous solution abiotic reduction of HgII by humic substances was highest in O2-free conditions and reduced in the presence of air. Obrist et al.54 showed that soil Hg0g emissions under dark laboratory conditions increased when upland soils changed from aerobic to oxygen-depleted, anaerobic conditions. Lalonde et al.83 calculated a redox potential of the HgI/Hg0aq couple of ∼1.3 V at pH 7 in water, a value that may decrease by several hundred millivolts if HgI is stabilized and thereby is in the range of potentials found in aerated soils (as high as +0.8 V vs NHE90). It is possible that mostly unsaturated soil moisture levels in the upland soils of our study generally favor oxidation leading to Hg0g immobilization, but that strong precipitation events and periods of saturation partially compensate for this through reduction and renewed Hg0g build-up in pores. Yet relationships to soil redox conditions likely are much more complex: in a wetlands study, Selvendiran et al. (2008) compared measurements at two sites and showed that at a meadow site, Hg0g fluxes initially showed atmospheric emissions during daytime and deposition during nighttime as well as following rain, but these switched to mainly emissions during the dryer summer. At a flooded riparian site with more

atmospheric concentrations due to the porous nature of the soil matrix and gas diffusion. How sinks or sources affect soil diffusion profiles in soils is well characterized for CO2. For example, strong concentration increases for CO2 with depth are the result of in-soil CO2 production, both by root respiration and microbial mineralization of litter and soil organic carbon.68 CO2 build-up in soil pores is then subject to diffusive transport within soils and into the atmosphere,68 following Fick’s first law of diffusion (as well as subject to other potential transport processes, such as wind pumping or advective transport). Hence, soil CO2 concentration profiles are affected both by the source strength of CO2 as well as soil physical properties (mainly diffusivity, which in soils is strongly controlled by porosity and water content). Two points are important to note: first, changes in soil diffusivity will lead to changes in the steepness and shape of CO2 concentration gradients,71 but will never change the direction of such gradients for a stable gas [this is different for nonstable (e.g., radioactive) gases such as 222Rn72]; second, observed gas concentrations do not directly represent the source strength at a depth, but rather the combined effects of sources and rates of diffusion (e.g., the highest CO2 concentrations are found in deep soils although highest CO2 production occurs in organic-rich top soils).72 Applying these concepts to observed pore Hg0g concentrations, we propose that concentrations below levels found in the atmosphere (which we term Hg0g immobilization from here forward) must be due to a net sink of Hg0g in mineral soils since concentrations otherwise would be in diffusive equilibrium with the atmosphere. Although the cycling and exchange of Hg0g between soil air and the soil matrix likely is characterized by a complex pattern of reduction, oxidation, and sorption/desorption processes (see the Introduction and ref 73), observed pore Hg0g immobilization provides evidence of a net Hg0g sink in mineral soils at both measurement sites and in all replicate soil profiles. Hg0g immobilization persisted throughout multiple seasons, and the strongest immobilization was observed in the uppermost horizons where we observed strongest concentration gradients (ΔC/Δz) averaging 11.4 ng m−3 m−1 and 15.4 ng m−3 m−1 (between the atmosphere and a soil depth of 7 cm) at the two sites, respectively. Lower concentration gradients were observed at lower depths, averaging 0.1 ng m−3 m−1 and 1.1 ng m−3 m−1, respectively, between 20 and 40 cm depths. While it is impossible to directly calculate a sink strength based on concentration measurements alone (as mentioned above; soil diffusivity was not measured), it is well-known that soils generally have the highest diffusivities in upper layers,72 therefore suggesting that the highest rates of Hg0g immobilization occurred within the uppermost 7 cm of the soil profile and lower rates at depths. Relationship of Pore Hg0g to Environmental Variables and Potential Mechanisms. On the basis of measurement of pore Hg0g concentration profiles alone, we cannot propose a mechanistic framework for the process leading to Hg0g immobilization in mineral soils. Several processes, however, may be involved based on the literature, including: adsorption of Hg0g to soil minerals or soil organic matter; abiotic or microbial oxidation of Hg0g to divalent Hg (HgII); or dissolution of Hg0g in soil solution. Sorption studies show that Hg0g sorbs well to soils, with increased sorption under high air Hg0g exposure.49,57,74−76 The sorption capacity in soils (e.g., estimated at 10 to 46 μg Hg0 kg−1 by Landa75) is orders of magnitude greater than the Hg0g sink estimated based on diffusion fluxes in our study (see 2246



Article

Across all measurements, however, it also seems that strongest pore Hg0g immobilization occurred not only under particularly dry (10 to 20% water content) but also when soils were very wet (35 to 45% water content), while being weakest under intermediate soil moisture conditions (25 to 35%; Figure 4B). The inverse correlation of pore Hg0g concentration to soil water content at low water content would not support that Hg0g concentrations were controlled mainly by dissolution in soil solution. However, we cannot eliminate the possibility that patterns shown in Figure 4B may be caused by different soil moisture responses of the two different sites, possibly due to different soil physical and chemical conditions (Table 1). An interesting observation is that across all measurements, we found significant linear regressions between pore Hg0g and CO2 concentrations (Figure 4A), which may show a link between Hg0g dynamics and microbial activity. Similarly, increased Hg0g concentrations after strong precipitation events were associated with increases in pore CO2 concentrations (Figure 3). Higher pore Hg0g concentrations under higher CO2 levelsalbeit still below atmospheric levelsmay be due to HgII reduction during microbial activity, which may partially compensate for pore Hg0g immobilization. Biological reduction of HgII to Hg0 by bacteria has been well documented,91,92 and involvement of microbes in Hg0g production has been reported in sediment and water studies.84,93−95 Evidence of involvement of microbial activity on soil Hg0g dynamics also has been reported based on changes in soil Hg0g emissions after manipulating microbial activity, including sterilization by autoclaving and methylbromine, addition of glucose, freeze−thawing cycles, and rewetting of dry soils.54,58,96,97 Laboratory studies performed in darkness also showed correlations of soil Hg0g emission and CO2 efflux rates,30,54,58 and this was observed in the field as well.98 However, the correlations between pore Hg0g and CO2 concentrations could also be coincidental, e.g., due to reduction in the transport of CO2 by lower soil diffusivity after rain while Hg0g concentrations declined due to increased dissolution in soil water. Comparison with Previous Studies and with Hg0g Fluxes Measured at the Surface. Two previous studies performed manual measurements of pore Hg0g concentrations in background soils55,64 and a few additional studies measured pore Hg0g concentrations in contaminated soils.65−67 Although studies show some differences with Hg0g patterns observed here, all measurements in background soils, in our assessment, support our observations of Hg0g immobilization in mineral

Figure 3. Effects of large precipitation events on soil pore Hg0g and CO2 concentrations. A. Concentrations of Hg0g and CO2 (insert) average for 10 days prior to and 10 days after the largest rainfall event (10.1 cm) at Blodgett Forest Research Station in October 2009. Second insert shows changes in soil moisture for the respective three soil depths of measurements. B. Same graph for a significant rainfall event (7.1 cm) at Sagehen Creek Field station in October 2010. Preand postprecipitation data for pore Hg0g and CO2 concentrations are statistically significant (P < 0.05) at all depths based on unpaired student t tests.

saturated conditions, however, fluxes were dominated by net Hg0g deposition. A second explanation of increased pore Hg0g concentrations after rain may be that precipitation provided fresh supply of labile Hg that percolated into the mineral soil where it may have been subject to reduction.

Figure 4. Relationships of soil pore Hg0g concentrations with CO2 concentrations and soil water content. A. Scatter plots and linear regressions between Hg0g concentrations and respective CO2 concentrations for observations in the top 7 cm layer (7 cm depth) in mineral soils. No consistent regressions were observed for deeper soil layers (20 and 40 cm). B. Scatter plots and linear regressions between Hg0g concentration and volumetric soil water content for the top 7 cm in mineral soils at the two sites. Linear regressions are statistically significant (P < 0.05). 2247



Article

layers.30,54 The notion that shallow zones of organic-rich surface soils and litter layers serve as Hg0g sources and overlay mineral soil layers where Hg0g immobilization dominates also are in support of Schlüter et al.17 who suggested that the organic litter horizons and organic-rich topsoils (A-horizons) are dominating sites for Hg0g evasion. In the absence of photochemical processes (i.e., at nights), however, upland soils therefore may become mainly Hg0g sinks. One study65 conducted in highly polluted soils concludes that an absence of clear pore Hg0g diffusion profiles in that study is evidence that surface Hg0g evasion is not dominated by diffusion processes, and that soil gas Hg0g flux is not diffusiondriven either. We do not agree, however, that the absence of Hg0g concentration gradients in that study is indicative that Hg0g diffusion is not an important process within soils. Soils are porous matrices, and gases are subject to diffusion according to Fick’s first law. As stated by ref 17, diffusion is the most probable mechanism for slow transport of gases across short distances. This does not mean that complex interactions, including sorption and desorption of Hg0 with organic matter or soil particles plus redox processes, are not also important in soils thus impacting diffusion profiles. In fact, we suggest that sorption, oxidation, or dissolution processes are likely responsible for observed pore Hg0g immobilization in mineral soil. Further, contaminated sites with pore air highly enriched in Hg0g,, e.g., reaching over 1000 ng m−3 (as in ref 65), may behave distinctly differently than background soils due to the presence of an Hg0g compensation point57,60,61 favoring Hg0g sorption and uptake. We suggest that all currently available pore Hg0g measurements in background soils are consistent with a Hg0g sink in mineral soils leading to pore Hg0g immobilization compared to atmosphere concentrations (our study) or concentrations in upper and near-surface organic-rich layers.55,64 Implications of Hg0g Immobilization in Mineral Soils for Soil and Global Cycling of Hg. An important question pertaining to an Hg0g sink in the mineral soil is whether gradients cause significant vertical diffusive fluxes and redistribute Hg0g within the soil profile, and possibly between soils and the atmosphere. To address this, we estimated the potential diffusive vertical soil flux, Fs (ng m−3 day−1), using concentration gradients between the top mineral soil layer and the atmosphere. This would provide a low estimate for vertical soil fluxes since upper soil Hg0g may be further enriched55,64 and since other transport processes such as advection are not included. We estimated the potential diffusive vertical Hg0g flux using Fick’s first law:

soils. SI Figure 3 shows box plots of measurements conducted at two sites in the Eastern U.S.55 where depth patterns can be described as follows: surface organic litter horizons and the interface of organic and mineral soil horizons show highest pore Hg0g concentrations, generally enhanced compared to atmospheric levels. These pore Hg0g enhancements therefore indicate surface or near-surface sources of Hg0g that then can emit to the atmosphere.99 Pore air in mineral soil layers, however, shows Hg0g levels that were lower compared to surface layers, and at times below levels in the atmosphere. Similar observations were made by Sigler and Lee64 at several sites in the northeastern U.S.; their measurements also show highest pore Hg0g concentrations (ranging from about 3.5 to 6 ng m−3; i.e., above atmospheric levels) near the organic-mineral soil horizon in support of a zone of Hg0g formation in the upper, organic soil and litter layers. In mineral soils at 20 and 50 cm depths, pore Hg0g levels were lower compared to the surface-zone, generally below 2 ng m−3, reaching as low as 0.3 ng m−3 at times. Therefore, both of these studies show a near-surface source of Hg0g; however, their observations of decreasing pore Hg0g concentrations in deeper layers also support zones of Hg0g sinks in mineral soils. Sigler and Lee 64 proposed that reasons for lower pore Hg0g concentrations at depth include a lower rate of Hg0g formation. Similar to CO2, however, any additional source of Hg0g in soils below the surface layer would further increase concentrations (i.e., equivalent to highest CO2 concentrations found in deeper soil layers despite a lower rate of production). Therefore, concentration declines of pore Hg0g at lower depths also point toward an Hg0g sink in mineral soils. In our study, pore Hg0g concentrations in the organic surface layers could not be measured because of the very thin organic surface layers in arid western forests which would have resulted in atmospheric advection during measurements. Consistent immobilization of Hg0g in soils, as observed in our study, suggesting an Hg0g sink seem to contradict measurements at the soil−atmosphere interface that show that Hg0g emission (i.e., an Hg0g source) is much more commonly observed than deposition.100,101 More recently, however, several surface flux studies reported that uncontaminated soils can be both sources and sinks for atmospheric Hg0g, such as in Adirondack forests;102 Eastern U.S. forests;34,35 Nevada soils;103,104 and other locations across the U.S.21,38,42,100,101,104,105 and Canada.100 Still, most studies report that both the frequency and magnitude of Hg0g emissions dominate over deposition, and that deposition is limited to nighttime periods or when surfaces are shaded. To reconcile the surface flux results with our observations, we propose that Hg0g emissions are primarily surface-driven, and are decoupled, from underlying soil Hg0g diffusion. We performed two sets of Hg0g exchange measurements using dynamic chambers at the soil surface (SI Table 2), and these measurements show net deposition of Hg0g to soils during nights but net Hg0g emission during sunlight exposure. These patterns agree well with strong diurnal Hg0g exchange fluxes measured over many soils related to photochemical processes.36−44,76 The lack of diurnal patterns in pore Hg0g concentrations are in support of the notion above. Surface Hg0g emissions therefore may be dominated by, or limited to, photochemical processes at the soil or litter surface, plus possibly light-independent Hg0g emissions from organic-rich upper soils and litter.55,64 In addition, carbon mineralization and litter decompositionwhich are strongest near the surfacemay contribute to Hg0g emissions in these upper

Fs = −Ds

ΔC Δz

where Ds is the Hg0g diffusion coefficient in the soil and ΔC/Δz is the concentration gradient between a soil depth z and the atmosphere.106 To estimate Ds, we used the approach proposed by Millington and Quirk 107 that employs soil porosity and soil water content to calculate gas diffusion coefficients in soils. Using this soil diffusion model, we estimated vertical downward fluxes by diffusion between the atmosphere and top mineral soil averaging 0.2 and 0.7 ng Hg m−2 day−1 for the two sites, respectivelyequivalent to 79 and 258 ng Hg m−2 y−1. Note that the presence of an organic surface layer with potentially enhanced Hg0g levels55,64 could increase diffusion fluxes between the upper soil zone and mineral soil layers, and at the same time also lead to efflux of Hg0g from the surface soil 2248



■

layer to the atmosphere. The main intent of these flux calculations, however, was to show that diffusive flux and vertical redistribution of Hg0g within the soil matrix was likely minor compared to other surface-atmosphere and in-soil processes. For example, wet Hg deposition ranges between 2.4 and 20.2 μg m−2 y−1, with an average deposition of 9.8 μg m−2 y−1 across 94 stations of the U.S. Mercury Deposition Network;108 and vertical percolation of Hg via soil water through the surface litter horizon into the mineral soil has been estimated at 33 μg m−2 y−1.109 Therefore, observed soil pore Hg0g immobilization unlikely leads to significant and noticeable accumulation of Hg in lower mineral soils. Therefore, the observations of an Hg0g sink in mineral soil layers are not in contradiction with observations that the highest Hg accumulation is observed in the upper, organic-rich soil layers.7−9 Another important implication of our results is that Hg is unlikely subject to Hg0g re-emission once incorporated into mineral soilsunless other processes such as runoff, erosion, solute transport, or disturbances remobilize Hg to the surface. Previous results showing correlations between soil CO2 and Hg0g evasion fluxes from soils31,98 as well as Hg0g losses during forest litter decomposition30 suggested losses of Hg0g upon terrestrial C mineralization. Results presented here suggest that such terrestrial Hg0g losses would be restricted to surface soil and litter layers, with no significant losses from mineral soils. This concept would agree with a recent study110 that suggested, based on abundance of naturally occurring stable isotopes, that the main source of Hg0g evading from soils is not legacy mercury from soils nor from wet deposition, but instead represents a more rapid cycling of recently atmospherically derived gas-phase Hg, possibly implying sorption of gas-phase mercury in upper soils and consecutive re-release in gaseous forms. A lack of, or reduction in, Hg0g re-emissions from mineral soils has strong implications for the global Hg cycle. Model simulations using the GEOS-Chem modela global 3D chemical transport model for atmospheric composition that tracks emission, deposition, and surface re-emission fluxes between the atmosphere and terrestrial and aquatic environmentsestimate global Hg0g emissions from terrestrial ecosystems ranging from 1300 to 2100 Mg y −1 and representing 18% to 28% of total atmospheric Hg emissions.23−25 Simulations incorporate mobilization and release of Hg0g from soils associated with decomposition of soil organic matter, accounting for about 900 Mg y−1 of terrestrial Hg0g emissions, with the remaining fractions caused by rapid photoreduction of HgII and surface revolatilization, from snow, ice, and vegetation.23 The predicted Hg0g mobilization associated with soil organic matter decomposition may be too high, however, if Hg0g re-emissions are limited to surface layers as our results suggest. A further implication of our findings is that soils play a key role for long-term Hg sequestration of past and current atmospheric Hg pollution with little gas-phase remobilization of Hg0g once incorporated into mineral soils. Large global terrestrial Hg reservoirs9,10 and the biogeochemical processes that remobilize these pools, therefore, are of key importance to understand terrestrial and atmospheric Hg loads. Current uncertainties associated with these processes highlight a need to better constrain Hg stored in soils and the processes that potentially re-emit and remobilize these Hg pools.

Article

ASSOCIATED CONTENT

S Supporting Information *

Detailed description and a figure (Figure S1) of the trace gas monitoring system to measure soil Hg0g and CO2 concentrations; additional information about measurement locations and auxiliary measurements such as soil water content, temperature, and soil physical and chemical properties; a table showing surface-flux results of Hg0g conducted with a dynamic chamber on top of the soil surface; a figure summarizing measurements of Hg0g and CO2 concentrations using box plots and providing statistics (Figure S2); and a figure summarizing previous soil Hg0g measurements by Moore and Castro (2012; Figure S3). This material is available free of charge via the Internet at http://pubs.acs.org.

■

AUTHOR INFORMATION

Corresponding Author

*Phone: (775) 674-7008; fax: (775) 673-7016; e-mail: daniel. [email protected]. Notes

The authors declare no competing financial interest.

■

ACKNOWLEDGMENTS This study was funded by the U.S. Environmental Protection Agency through a Science-To-Achieve-results grant (No. R833378) and received supplementary support through a DRI Internal Project Assignment grant and an NSF-funded project (Award No. 1313755CNH: Managing Impacts of Global Transport of Atmosphere-Surface Exchangeable Pollutants in the Context of Global Change). We thank C. Berger and L. Arnone for help with data processing, and R. Kreidberg with editorial support.

■

REFERENCES

(1) 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), 1−7. (2) Lorey, P.; Driscoll, C. T. Historical trends of mercury deposition in Adirondack lakes. Environ. Sci. Technol. 1999, 33 (5), 718−722. (3) 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. (4) Teisserenc, R.; Lucotte, M.; Houel, S. Terrestrial organic matter biomarkers as tracers of Hg sources in lake sediments. Biogeochemistry 2011, 103 (1−3), 235−244. (5) 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. (6) 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. (7) Obrist, D.; Johnson, D. W.; Lindberg, S. E.; Luo, Y.; Hararuk, O.; Bracho, R.; Battles, J. J.; Dail, D. B.; Edmonds, R. L.; Monson, R. K.; Ollinger, S. V.; Pallardy, S. G.; Pregitzer, K. S.; Todd, D. E. Mercury distribution across 14 U.S. forests. Part I: Spatial patterns of concentrations in biomass, litter, and soils. Environ. Sci. Technol. 2011, 3974−3981. (8) Grigal, D. F. Mercury sequestration in forests and peatlands: A review. J. Environ. Qual. 2003, 32 (2), 393−405. (9) Obrist, D. Mercury distribution across 14 U.S. forests. Part II: Patterns of methyl mercury concentrations and areal mass of total and methyl mercury. Environ. Sci. Technol. 2012, 46, 5921−5930.

2249



Article

(10) Hararuk, O.; Obrist, D.; Luo, Y. Modeling the sensitivity of soil mercury storage to climate-induced changes in soil carbon pools. Biogeosciences 2013, 10, 2393−2407. (11) Lag, J.; Steinnes, E. Regional distribution of mercury in humus layers of Norwegian forest soils. Act. Agr. Scand. 1978, 28 (4), 393− 396. (12) 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. (13) 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. (14) Lindqvist, O.; Johansson, K.; Aastrup, M.; Anderson, A.; Bringmark, L.; Hovenius, G.; Hakanson, L.; Iverfeldt, A.; Meili, M.; Timm, B. Mercury in the Swedish environmentRecent research on causes, consequences and corrective methods. Water Air Soil Pollut. 1991, 55, 1−261. (15) 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. (16) Schuster, E. The behavior of mercury in the soil with special emphasis on complexation and adsorption processesA review of the literature. Water Air Soil Pollut. 1991, 56, 667−680. (17) Schluter, K. Review: Evaporation of mercury from soils. An integration and synthesis of current knowledge. Environ. Geol. 2000, 39 (3−4), 249−271. (18) Yin, Y. J.; Allen, H. E.; Li, Y. M.; Huang, C. P.; Sanders, P. F. Adsorption of mercury(II) by soil: Effects of pH, chloride, and organic matter. J. Environ. Qual. 1996, 25 (4), 837−844. (19) 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. (20) 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 (3), 855−865. (21) Gustin, M. S.; Lindberg, S. E.; Weisberg, P. J. An update on the natural sources and sinks of atmospheric mercury. Appl. Geochem. 2008, 23 (3), 482−493. (22) 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. Cyc. 2008, 22, (3). (23) Smith-Downey, N. V.; Sunderland, E. M.; Jacob, D. J. Anthropogenic impacts on global storage and emissions of mercury from terrestrial soils: Insights from a new global model. J. Geophys. Res.-Biogeosci. 2010, 115, 11. (24) Corbitt, E. S.; Jacob, D. J.; Holmes, C. D.; Streets, D. G.; Sunderland, E. M. Global source-receptor relationships for mercury deposition under present-day and 2050 emissions scenarios. Environ. Sci. Technol. 2011, 45 (24), 10477−10484. (25) Amos, H. M.; Jacob, D. J.; Streets, D. G.; Sunderland, E. M. Legacy impacts of all-time anthropogenic emissions on the global mercury cycle. Global Biogeochem. Cyc. 2013, DOI: 10.1002/ gbc.20040. (26) Streets, D. G.; Devane, M. K.; Lu, Z.; Bond, T. C.; Sunderland, E. M.; Jacob, D. J. All-time releases of mercury to the atmosphere from human activities. Environ. Sci. Technol. 2011, 45 (24), 10485−10491. (27) Lindberg, S.; Bullock, R.; Ebinghaus, R.; Engstrom, D.; Feng, X. B.; Fitzgerald, W.; Pirrone, N.; Prestbo, E.; Seigneur, C. A synthesis of progress and uncertainties in attributing the sources of mercury in deposition. Ambio 2007, 36 (1), 19−32. (28) Ericksen, J. A.; Gustin, M. S.; Lindberg, S. E.; Olund, S. E.; 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.

(29) Harris, R. C.; Rudd, J. W. M.; Amyot, 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.; Southworth, G. R.; St. Louis, V. L.; 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. (30) Pokharel, A. K.; Obrist, D. Fate of mercury in tree litter during decomposition. Biogeosciences 2011, 8 (9), 2507−2521. (31) Obrist, D.; Faïn, X.; Berger, C. Gaseous elemental mercury emissions and CO2 respiration rates in terrestrial soils under controlled aerobic and anaerobic laboratory conditions. Sci. Total Environ. 2010, 408 (7), 1691−1700. (32) 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 (2−3), 851− 863. (33) 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. (34) Kuiken, T.; Gustin, M.; Zhang, H.; Lindberg, S.; Sedinger, B. Mercury emission from terrestrial background surfaces in the eastern USA. II: Air/surface exchange of mercury within forests from South Carolina to New England. Appl. Geochem. 2008, 23 (3), 356−368. (35) Kuiken, T.; Zhang, H.; Gustin, M.; Lindberg, S. Mercury emission from terrestrial background surfaces in the eastern USA. Part I: Air/surface exchange of mercury within a southeastern deciduous forest (Tennessee) over one year. Appl. Geochem. 2008, 23 (3), 345− 355. (36) Carpi, A.; Lindberg, S. E. Sunlight-mediated emission of elemental mercury from soil amended with municipal sewage sludge. Environ. Sci. Technol. 1997, 31 (7), 2085−2091. (37) Moore, C.; Carpi, A. Mechanisms of the emission of mercury from soil: Role of UV radiation. J. Geophys. Res.-Atmos. 2005, 110 (D24), 9. (38) Poissant, L.; Casimir, A. Water−air and soil−air exchange rate of total gaseous mercury measured at background sites. Atmos. Environ. 1998, 32 (5), 883−893. (39) Engle, M. A.; Gustin, M. S.; Zhang, H. Quantifying natural source mercury emissions from the Ivanhoe Mining District, northcentral Nevada, USA. Atmos. Environ. 2001, 35 (23), 3987−3997. (40) Zhang, H.; Lindberg, S. E. Sunlight and Iron(III)-induced photochemical production of dissolved gaseous mercury in freshwater. Environ. Sci. Technol. 2001, 35, 928−935. (41) Gustin, M. S.; Taylor, G. E.; Maxey, R. A. Effect of temperature and air movement on the flux of elemental mercury from substrate to the atmosphere. J. Geophys. Res.-Atmos. 1997, 102 (D3), 3891−3898. (42) Zhang, H.; Lindberg, S. E.; Marsik, F. J.; Keeler, G. J. Mercury air/surface exchange kinetics of background soils of the Tahquamenon River watershed in the Michigan Upper Peninsula. Water Air Soil Pollut. 2001, 126 (1−2), 151−169. (43) Xiao, Z. F.; Munthe, J.; Schroeder, W. H.; Lindqvist, O. Vertical fluxes of volatile mercury over forest soil and lake surfaces in Sweden. Tellus Ser. B-Chem. Phys. Meteorol. 1991, 43 (3), 267−279. (44) Bahlmann, E.; Ebinghaus, R. In Process Studies on Mercury Fluxes over Different Soils with a Laboratory Flux Measurement System (LFMS), XIIth International Conference on Heavy Metals in the Environment, Grenoble, France, 2003; Boutron, C., Ferrari, C., Eds.; EDP Sciences: Grenoble, France, 2003; pp 99−102. (45) Lindberg, S. E.; Zhang, H.; Gustin, M.; Vette, A.; Marsik, F.; Owens, J.; Casimir, A.; Ebinghaus, R.; Edwards, G.; Fitzgerald, C.; Kemp, J.; Kock, H. H.; London, J.; Majewski, M.; Poissant, L.; Pilote, M.; Rasmussen, P.; Schaedlich, F.; Schneeberger, D.; Sommar, J.; Turner, R.; Wallschlager, D.; Xiao, Z. Increases in mercury emissions from desert soils in response to rainfall and irrigation. J. Geophys. Res.Atmos. 1999, 104 (D17), 21879−21888. 2250



Article

control of Hg evasion from soils. Sci. Total Environ. 2003, 304 (1−3), 175−184. (66) Kromer, E.; Friedrich, G.; Wallner, P. Mercury and mercurycompounds in the surface air, soil gas, soils and rocks. J. Geochem. Explor. 1981, 15 (1−3), 51−62. (67) Wallschläger, D.; Kock, H. H.; Schroeder, W. H.; Lindberg, S. E.; Ebinghaus, R.; Wilken, R. D. Estimating gaseous mercury emissions from contaminated floodplain soils to the atmosphere with simple field measurement techniques. Water Air Soil Pollut. 2002, 135 (1−4), 39− 54. (68) Rolston, D. E. Gas Diffusivity; American Society of Agronomy, Inc. Soil Science Society of America, Inc.: Madison, WI, 1986; Vol. 5. (69) Buchmann, N.; Kao, W. Y.; Ehleringer, J. R. Carbon dioxide concentrations within forest canopies - Variation with time, stand structure, and vegetation type. Glob. Change Biol. 1996, 2 (5), 421− 432. (70) Sprovieri, F.; Pirrone, N.; Ebinghaus, R.; Kock, H.; Dommergue, A. A review of worldwide atmospheric mercury measurements (vol 10, pg 8245, 2010). Atmos. Chem. Phys. 2010, 10 (17), 8531−8531. (71) Koehler, B.; Zehe, E.; Corre, M. D.; Veldkamp, E. An inverse analysis reveals limitations of the soil-CO2 profile method to calculate CO2 production and efflux for well-structured soils. Biogeosciences 2010, 7 (8), 2311−2325. (72) Davidson, E. A.; Trumbore, S. E. Gas diffusivity and production of CO2 in deep soils of the eastern Amazon. Tellus Ser. B-Chem. Phys. Meteorol. 1995, 47 (5), 550−565. (73) Zhang, H.; Lindberg, S. E. Processes influencing the emission of mercury from soils: A conceptual model. J. Geophys. Res.-Atmos. 1999, 104 (D17), 21889−21896. (74) Fang, S. C. Sorption and transformation of mercury-vapor by dry soils. Environ. Sci. Technol. 1978, 12 (3), 285−288. (75) Landa, E. R. Retention of metallic mercury-vapor by soils. Geochim. Cosmochim. Acta 1978, 42 (9), 1407−1411. (76) Xin, M.; Gustin, M.; Johnson, D. Laboratory investigation of the potential for re-emission of atmospherically derived Hg from soils. Environ. Sci. Technol. 2007, 41 (14), 4946−4951. (77) do Valle, C. M.; Santana, G. P.; Windmoller, C. C. Mercury conversion processes in Amazon soils evaluated by thermodesorption analysis. Chemosphere 2006, 65 (11), 1966−1975. (78) Amyot, M.; Gill, G. A.; Morel, F. M. M. Production and loss of dissolved gaseous mercury in coastal seawater. Environ. Sci. Technol. 1997, 31 (12), 3606−3611. (79) Lalonde, J. D.; Amyot, M.; Kraepiel, A. M. L.; Morel, F. M. M. Photooxidation of Hg(0) in artificial and natural waters. Environ. Sci. Technol. 2001, 35 (7), 1367−1372. (80) Amyot, M.; Morel, F. M. M.; Ariya, P. A. Dark oxidation of dissolved and liquid elemental mercury in aquatic environments. Environ. Sci. Technol. 2005, 39 (1), 110−114. (81) Amyot, M.; Mierle, G.; Lean, D.; McQueen, D. J. Effect of solar radiation on the formation of dissolved gaseous mercury in temperate lakes. Geochim. Cosmochim. Acta 1997, 61 (5), 975−987. (82) Mason, R. P.; Lawson, N. M.; Sheu, G. R. Mercury in the Atlantic Ocean: Factors controlling air−sea exchange of mercury and its distribution in the upper waters. Deep-Sea Res. Part II-Top. Stud. Oceanogr. 2001, 48 (13), 2829−2853. (83) Lalonde, J. D.; Amyot, M.; Orvoine, J.; Morel, F. M. M.; Auclair, J. C.; Ariya, P. A. Photoinduced oxidation of Hg-0 (aq) in the waters from the St. Lawrence estuary. Environ. Sci. Technol. 2004, 38 (2), 508−514. (84) Whalin, L.; Kim, E. H.; Mason, R. Factors influencing the oxidation, reduction, methylation and demethylation of mercury species in coastal waters. Mar. Chem. 2007, 107 (3), 278−294. (85) Yamamoto, M. Stimulation of elemental mercury oxidation in the presence of chloride ion in aquatic environments. Chemosphere 1996, 32 (6), 1217−1224. (86) Hines, N. A.; Brezonik, P. L. Mercury dynamics in a small Northern Minnesota lake: Water to air exchange and photoreactions of mercury. Mar. Chem. 2004, 90 (1−4), 137−149.

(46) Gustin, M.; Ericksen, J.; Schorran, D.; Johnson, D.; Lindberg, S.; Coleman, J. Application of controlled mesocosms for understanding mercury air−soil−plant exchange. Environ. Sci. Technol. 2004, 38 (22), 6044−6050. (47) Gustin, M. S.; Stamenkovic, J. Effect of watering and soil moisture on mercury emissions from soils. Biogeochemistry 2005, 76 (2), 215−232. (48) Poissant, L.; Pilote, M.; Casimir, A. Mercury flux measurements in a naturally enriched area: Correlation with environmental conditions during the Nevada Study and Tests of the Release of Mercury From Soils (STORMS). J. Geophys. Res.-Atmos. 1999, 104 (D17), 21845−21857. (49) Fang, S. C. Studies on the sorption of elemental mercury-vapor in soils. Arch. Environ. Contam. Toxicol. 1981, 10 (2), 193−201. (50) Song, X. X.; Van Heyst, B. Volatilization of mercury from soils in response to simulated precipitation. Atmos. Environ. 2005, 39 (39), 7494−7505. (51) Gustin, M. S.; Lindberg, S.; Marsik, F.; Casimir, A.; Ebinghaus, R.; Edwards, G.; Hubble-Fitzgerald, C.; Kemp, R.; Kock, H.; Leonard, T.; London, J.; Majewski, M.; Owens, J.; Pilote, M.; Poissant, L.; Rasmussen, P.; Schaedlich, F.; Schneeberger, D.; Schroeder, W.; Sommar, J.; Turner, R.; Vette, A.; Wallschlaeger, D.; Xiao, Z.; Zhang, H. Nevada SToRMS Project: Measurement of mercury emissions from naturally enriched surfaces. J. Geophys. Res. 1999, 104, 21831−21844. (52) Gustin, M. S.; Lindberg, S. E.; Kenneth, A.; Coolbaugh, M.; lan Vette, A.; Zhang, H., Assessing the contribution of natural sources to regional atmospheric mercury budgets. Sci. Total Environ. 2000, 259, (61−71). (53) Engle, M.; Gustin, M.; Zhang, H. Quantifying natural source mercury emissions from the Ivanhoe Mining District, north-central Nevada, USA. Atmos. Environ. 2001, 35 (23), 3987−3997. (54) 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. (55) Moore, C. W.; Castro, M. S. Investigation of factors affecting gaseous mercury concentrations in soils. Sci. Total Environ. 2012, 419 (0), 136−143. (56) Mauclair, C.; Layshock, J.; Carpi, A. Quantifying the effect of humic matter on the emission of mercury from artificial soil surfaces. Appl. Geochem. 2008, 23 (3), 594−601. (57) Xin, M.; Gustin, M. S. Gaseous elemental mercury exchange with low mercury containing soils: Investigation of controlling factors. Appl. Geochem. 2007, 22, 1451−1466. (58) Fritsche, J.; Obrist, D.; Alewell, C. Evidence of microbial control of Hg-0 emissions from uncontaminated terrestrial soils. J. Plant Nutr. Soil Sci. 2008, 171 (2), 200−209. (59) Obrist, D.; Fain, X.; Berger, C. Gaseous elemental mercury emissions and CO2 respiration rates in terrestrial soils under controlled aerobic and anaerobic laboratory conditions. Sci. Total Environ. 2010, 408 (7), 1691−1700. (60) Hanson, P. J.; Lindberg, S.; Tabberer, T.; Owens, J.; Kim, K.-H. Foliar exchange of mercury vapor: evidence for a compensation point. Water Air Soil Pollut. 1995, 80, 373−382. (61) Ericksen, J. A.; Gustin, M. S. Foliar exchange of mercury as a function of soil and air mercury concentrations. Sci. Total Environ. 2004, 324 (1−3), 271−279. (62) Engle, M. A.; Gustin, M. S.; Lindberg, S. E.; Gertler, A. W.; Ariya, P. A. The influence of ozone on atmospheric emissions of gaseous elemental mercury and reactive gaseous mercury from substrates. Atmos. Environ. 2005, 39 (39), 7506−7517. (63) Zhang, H.; Lindberg, S. E.; Kuiken, T. Mysterious diel cycles of mercury emission from soils held in the dark at constant temperature. Atmos. Environ. 2008, 42 (21), 5424−5433. (64) Sigler, J. M.; Lee, X. Gaseous mercury in background forest soil in the northeastern United States. J. Geophys. Res.-Biogeosci. 2006, 111, (G2). (65) Johnson, D. W.; Benesch, J. A.; Gustin, M. S.; Schorran, D. S.; Lindberg, S. E.; Coleman, J. S. Experimental evidence against diffusion 2251



Article

and the global mercury cycle. Global Biogeochem. Cyc. 2013, 27 (1), 222−238.

(87) Schroeder, W. H.; Munthe, J. Atmospheric mercuryAn overview. Atmos. Environ. 1998, 32 (5), 809−822. (88) Hebert, V. R.; Miller, G. C. Depth dependence of direct and indirect photolysis on soil surfaces. J. Agric. Food Chem. 1990, 38 (3), 913−918. (89) Allard, B.; Arsenie, I. Abiotic reduction of mercury by humic substances in aquatic systemAn important process for the mercury cycle. Water Air Soil Pollut. 1991, 56, 457−464. (90) Scheffer, F.; Schachtschabel, P., Lehrbuch der Bodenkunde, 15th ed; Spektrum Akademischer Verlag GmbH: Heidelberg, Berlin, Germany, 2002; pp 136−142. (91) Hansen, C. L.; Zwolinski, G.; Martin, D.; Williams, J. W. Bacterial removal of mercury from sewage. Biotechnol. Bioeng. 1984, 26 (11), 1330−1333. (92) Lovley, D. R. Dissimilaroty metal reduction. Annu. Rev. Microbiol. 1993, 47, 263−290. (93) Mason, R. P.; Morel, F. M. M.; Hemond, H. F. The role of microorganisms in elemental mercury formation in natural waters. Water Air Soil Pollut. 1995, 80 (1−4), 775−787. (94) Siciliano, S. D.; O’Driscoll, N. J.; Lean, D. R. S. Microbial reduction and oxidation of mercury in freshwater lakes. Environ. Sci. Technol. 2002, 36 (14), 3064−3068. (95) Rolfhus, K. R.; Fitzgerald, W. F. Mechanisms and temporal variability of dissolved gaseous mercury production in coastal seawater. Mar. Chem. 2004, 90 (1−4), 125−136. (96) Landa, E. R. Microbial aspects of volatile loss of applied mercury(II) from soils. J. Environ. Qual. 1978, 7 (1), 84−86. (97) Rogers, R. D.; McFarlane, J. C. Factors influencing the volatilization of mercury from soils. J. Environ. Qual. 1979, 8 (2), 255− 260. (98) Wickland, K.; Krabenhoft, D.; Olund, S. Evidence for a Link between Soil Respiration and Mercury Emissions from Organic Soils. In 8th International Conference on Mercury as a Global Pollutant; Madison, WI, 2006. (99) Moore, C. W. Factors Influencing Surface Atmosphere Exchange of Gaseous Elemental Mercury in Western Maryland; University of Maryland: College Park, MD, 2011. (100) Schroeder, W.; Beauchamp, S.; Edwards, G.; Poissant, L.; Rasmussen, P.; Tordon, R.; Dias, G.; Kemp, J.; Van Heyst, B.; Banic, C. Gaseous mercury emission from natural sources in Canadian landscapes. J. Geophys. Res.-Atmos. 2005, 110. (101) 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. (102) Choi, H.-D.; Holsen, T. M. Gaseous mercury fluxes from the forest floor of the Adirondacks. Environ. Pollut. 2009, 157 (2), 592− 600. (103) Gustin, M. S.; Engle, M.; Ericksen, J.; Lyman, S.; Stamenkovic, J.; Xin, M. Mercury exchange between the atmosphere and low mercury containing substrates. Appl. Geochem. 2006, 21 (11), 1913− 1923. (104) 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. (105) Nacht, D. M.; Gustin, M. S. Mercury emissions from background and altered geologic units throughout Nevada. Water Air Soil Pollut. 2004, 151 (1−4), 179−193. (106) Rolston, D. E. Gas Flux; American Society of Agronomy, Inc./ Soil Science Society of America, Inc.: Madison, WI, 1986; Vol. 5. (107) Millington, R. J.; Quirk, J. P. Permeability of porous solids. Trans. Faraday Soc. 1961, 57, 1200−1207. (108) NADP. The Mercury Deposition Network. http://nadp.sws. uiuc.edu/MDN/. (109) 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. (110) Demers, J. D.; Blum, J. D.; Zak, D. R. Mercury isotopes in a forested ecosystem: Implications for air−surface exchange dynamics 2252


Vertical Profile Measurements of Soil Air Suggest Immobilization of

Recommend Documents