Chemosphere 134 (2015) 113–119
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Inhibition of mercury release from forest soil by high atmospheric deposition of Ca2+ and SO2 4 Yao Luo a, Lei Duan a,b,⇑, Guangyi Xu c, Jiming Hao a,b a
State Key Laboratory of Environmental Simulation and Pollution Control, School of Environment, Tsinghua University, Beijing 100084, China Collaborative Innovation Centre for Regional Environmental Quality, Tsinghua University, Beijing 100084, China c Shenzhen Academy of Environment Sciences, Shenzhen 518001, China b
h i g h l i g h t s 2+
Hg release from forest soil was inhibited by high deposition of Ca
and SO2 4 .
The elevated organic S enhanced the capacity of surface soil to bind Hg. Soil Hg emission may increase in the future due to stringent control of PM and SO2.
a r t i c l e
i n f o
Article history: Received 1 July 2014 Received in revised form 14 March 2015 Accepted 23 March 2015 Available online 15 May 2015 Handling Editor: X. Cao Keywords: Flue gas desulfurization gypsum (FGDG) Mercury Natural source Forest soil Organic sulfur Atmospheric deposition
a b s t r a c t As one of the most important natural mercury (Hg) sources, soil release (emission to the atmosphere or leaching to soil water) depends on various factors, some of which can be affected by atmospheric deposition. We studied the effect of flue gas desulfurization gypsum (FGDG) addition on soil Hg release in a Masson pine (Pinus massoniana) forest in southwestern China. FGDG addition simulated atmospheric deposition of Ca2+, SO2 4 and Hg, which are commonly high in China. Results showed that Hg concentration in soil water decreased with the gypsum treatment, suggesting that the mobility of Hg in mineral soil was reduced. Moreover, the application of gypsum also seems to have decreased Hg emission from the soil, shown by the lower Hg contents in leaf tissues of ground vegetation in the treated plots than in the reference. Both Hg mobility in the soil and Hg emission to the atmosphere were decreased despite the additional Hg input from FGDG. The decreased DOC concentration in soil water and the elevated organic sulfur content in the soil Oe & Oa horizons were speculated to result in an enhanced capacity of surface soil to bind Hg, and thus to reduce Hg release from the soil. However, with the increasingly stringent control of particulate matter (PM) and sulfur dioxide (SO2) emissions in China, the deposition of Ca2+ and SO2 4 is expected to decrease, and their ability to inhibit soil Hg release is likely to decline in the future. Ó 2015 Elsevier Ltd. All rights reserved.
1. Introduction Mercury pollution is a global environmental concern due to its harm to human beings (Jensen and Jernelöv, 1969; Mergler et al., 2007). In terrestrial ecosystem, forest soil is an important natural source of Hg. Not only can Hg be leached out from forest soil through soil water (Hintelmann et al., 2002), but also large amounts of Hg emit to the atmosphere (Lindberg et al., 1979; Xiao et al., 1991). The soil properties, such as pH value, Hg content ⇑ Corresponding author at: State Key laboratory of Environmental Simulation and Pollution Control, School of Environment, Tsinghua University, Beijing 100084, China. Tel.: +86 (10)62783758; fax: +86 (10)62773957. E-mail address: [email protected] (L. Duan). http://dx.doi.org/10.1016/j.chemosphere.2015.03.081 0045-6535/Ó 2015 Elsevier Ltd. All rights reserved.
and humic matter content, may influence Hg release significantly (Gillis and Miller, 2000; Mauclair et al., 2008). As soil properties can be greatly affected by atmospheric deposition (e.g., soil acidification), it is important to clarify the effects of atmospheric deposition on forest soil Hg release. China is currently the world’s top emitter of anthropogenic Hg due to enormous coal combustion, nonferrous metals production and so on (Pirrone et al., 2010; Wang et al., 2010; Wu et al., 2012), which results in high Hg deposition in China. The Hg deposition in remote areas and urban areas in China were 1–2 times and 1–2 magnitude respectively higher than those in North America and Europe (Fu et al., 2012). Recent estimation using CMAQ-Hg model showed that dry deposition of Hg in the east of China in 2005 was generally more than 40 lg m2 yr1, and wet deposition
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more than 20 lg m2 yr1 (Lin et al., 2010), much higher than the total Hg deposition in northeastern North America (21.3 lg m2 yr1; Miller et al., 2005) and in Japan (20.8 lg m2 yr1; Sakata and Marumoto, 2005). Large amounts of Hg in soil deposited from atmosphere would be harmful to human beings or reemitted from soil. Due to serious pollution of particulate matter (PM) and sulfur dioxide (SO2), atmospheric deposition in China has also high con-
1971 to 2000 (Meteorological Station of Chongqing), with an annual average precipitation pH of 4.0–4.2 (Larssen et al., 2011). Atmospheric deposition of sulfur (8–10 keq ha1 yr1), nitrogen (about 3 keq ha1 yr1, with two-thirds NH+4 and one-third NO 3) and Ca2+ (about 6 keq ha1 yr1) is high and the soil is a Haplic Acrisol which is acidic (pH 3.6–4.2) (Larssen et al., 2011).
2+ tents of Ca2+ and SO2 and SO2 4 . The deposition fluxes of Ca 4 at some monitored sites in southern and southwestern China (Ca2+:
2.2. FGDG component analysis and application
1 0.8–5.7 keq ha1 yr1, SO2 yr1; Larssen et al., 4 : 1.5–10.5 keq ha 2011) are much higher than those in North America and Europe during the peak period of acid deposition. This high deposition may change soil chemical properties and then influence the soil Hg release. With the increasingly stringent control of PM and SO2
The gypsum (FGDG) used in this study was obtained from a coal-fired power plant in Chongqing city. Total Hg content of the gypsum was analyzed by a direct Hg analyzer (DMA80, Milestone Inc., Italy), and Soluble Hg content was measured by sulfuric acid leaching (see SI). The contents of other elements (e.g. Ca and S) were measured by ICP-AES (Thermo Jarrell Ash, USA) after being decomposed by concentrated nitric acid. Carbonate (CaCO3) content was quantified using a titration method. The gypsum (FGDG) used in this study was about 88% pure (CaSO42H2O), with 8% of CaCO3 due to the use of excessive desulfurizing agent (limestone) in the FGD process. The amount of CaCO3 addition was approximately 1.00 keq ha1 yr1, which could neutralize about 10% of sulfur deposition. The total Hg concentration was 910 ± 10 lg kg1, and around 20% was soluble Hg. The amount of soluble Hg did not vary with water pH according to the leaching experiments (Table S1 in SI), indicating that the insoluble Hg in FGDG was quite stable against acid and might be mercuric sulfide (HgS) (Rallo et al., 2010). In October 2009, six 10 10 m2 plots were established in a homogeneous forest. The distance between the nearest plots was about 10 m. The plots were randomly divided into two groups, each with three replicates. One group was treated by the gypsum (FGDG) and the other was used as a control, denoted as FGDG and Control, respectively. The gypsum was sprinkled onto the forest floor manually at a dose of 63.4 g m2 in December 2009 and then at a dose of 70.0 g m2 in May 2011 for each FGDG plot. The dose was chosen to approximately double the annual deposition of Ca2+ in this area. Buffer zone with 1 m width around each FGDG plot was also treated. The visible layer of the gypsum disappeared soon after rain and did not influence the air/surface exchange.
emissions in China, however, the deposition of Ca2+ and SO2 4 and its influence on the Hg release are expected to decrease in the future (Zhao et al., 2011). In order to study the effects of the high deposition of Ca2+ and SO2 4 on soil Hg release, we applied flue gas desulfurization gypsum (FGDG), which is generated in the process of flue gas desulfurization (FGD) and mainly consists of CaSO4 and a little residual CaCO3, on forest soil to simulate the additional atmospheric deposition of Ca2+, SO2 and Hg. Gypsum (CaSO42H2O) was used 4 solution because more than two thirds instead of Ca2+ and SO2 4 of the deposition of Ca2+ and SO2 4 in China is through dry deposiin atmotion (Larssen et al., 2011), and most of Ca2+ and SO2 4 spheric particles exists as gypsum (Takahashi et al., 2008; Quan et al., 2008; Ma et al., 2013). The gypsum also contains some Hg, because Hg in flue gas can be transferred into the gypsum during the FGD process (Kairies et al., 2006; Wang et al., 2010). After the gypsum application, the decrease in Hg concentration in soil water was expected to be found, because the gypsum could decrease DOC concentration in soil water which showed strong positive correlation to Hg concentration in water (Evans et al., 1988; Akerblom et al., 2008). However, the change of soil Hg0 emission was uncertain. On one side, Hg0 emission may be increased because of the additional Hg inputs (Coolbaugh et al., 2002); on the other side, gypsum could elevate organic nitrogen content in soil (Belkacem and Nys, 1995) which could enhance the capability of soil to bind Hg and decrease Hg0 emission. Hg leaching in soil water and Hg content in soil and vegetation in the forest were investigated during 2010–2011. Although Hg emission flux from soil was not measured directly, the change of flux was reflected by the change of Hg content in leaf tissues of ground vegetation, because numerous studies have indicated that most of the Hg in leaves originates from the atmosphere (Ericksen et al., 2003; Ericksen and Gustin, 2004; Millhollen et al., 2006; Fay and Gustin, 2007). Here we hypothesize: (1) Hg concentration in soil water will be decreased by the gypsum application; and (2) the soil Hg0 emission will be influenced by the changes of soil chemical properties, which can be reflected in leaves tissues. 2. Materials and methods 2.1. Site description The field manipulation experiment was carried out in a Masson pine (Pinus massoniana) forest at Tieshanping (29°380 N, 106°410 E), about 25 km northeast of Chongqing City in southwestern China. Masson pine forests, either natural or planted, are widely distributed in the humid subtropical areas of southern China (Wu, 1980). The annual mean temperature and precipitation at the experimental site were 18.2 °C and 1105 mm respectively from
2.3. Sample collection and chemical analysis In October 2009, four throughfall (TF) collectors were placed in the corners of each plot, and two series of lysimeters were installed close to the center of each plot. The throughfall collector consisted of a funnel with a certain sectional area to collect throughfall and a bottle to store water. The material of the two sections was PET (polyethylene terephthalate), which was found to be hard to absorb mercury (Parker and Bloom, 2005). Each series of lysimeters contained a zero-tension plate lysimeters (0.30 0.30 m2) installed beneath the organic layer (S0), and three rhizons (which are made of inert polymer with very low metal absorption and suitable for trace metal research; Knight et al., 1998) installed at 5, 15, and 30 cm depth in the mineral soil (layers S1, S2, and S3). One series was used for Hg measurement, the sample bottles of which were pre-added with 4 mL of 6 mol L1-ultrapure HCl to inhibit the transformation and absorption of Hg in the bottles. The other series was used for measurements of pH value, ion concentrations and dissolved organic carbon (DOC). Since the beginning of 2010, throughfall and soil water from each layer were gathered weekly and refrigerated immediately. The samples of every four weeks were pooled into a bulk sample (monthly sample) for chemical analysis.
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Litterfall in each plot was gathered monthly from four 1 1 m2 square collectors, each close to a throughfall collector. Herbaceous ground vegetation was examined in five permanently marked 1 1 m2 quadrates in each plot. The five quadrates were distributed in such a way that each quadrat contained at least one of the major species in the forest (Woodwardia japonica, Dryopteris fuscipes, Dicranopteris pedata or Miscanthus sinensis), and all major species were included in the five quadrates. Only current-year leaves of those four species, which is generally 30–60 cm in height, were collected from each quadrate for chemical analysis. For fern (W. japonica, D. fuscipes and D. pedata), three pinnately compound leaves were completely cut by a steel scissor; for grass (M. sinensis), at least five leaves were cut. To measure needle Hg content, three branches were cut randomly from the crown top of each of the two highest Masson pines in each plot (15–25 m), and 1000 current-year needles were gathered from the branches. The leaves sampling was conducted in early November (at the end of the growing season) of each year. Soil was collected in October 2009 (before treatment) and October 2011 (about two years after treatment) from six layers of each plot, i.e., Oi horizon (undecomposed litter layer), Oe & Oa horizons (moderately decomposed organic matter layer & humus layer), A horizon (mineral soil 0–2 cm) and B1–B3 horizons (mineral soil 2–5 cm, 5–15 cm, and 15–30 cm, respectively). In each plot, three soil profiles were sampled, and three samples from the same layer were mixed to one sample. Organic soil was picked from forest floor using clean latex gloves and stainless spade, while mineral soil was collected using soil auger. All soil and plant samples were immediately bagged in plastic bags and stored in an insulation box with ice bags until transport to laboratory. After pH was analyzed, water samples were filtered through a 0.45-lm glass fiber filter. Total dissolved Hg concentration was measured by cold vapor atomic fluorescence spectrometry (CVAFS) (Tekran model 2600 system, Tekran Inc., Canada), following the modified U.S. EPA Method 1631 guideline (Holmes and þ Lean, 2006). Concentrations of other ions (e.g. NO 3 and NH4 ) were measured by ion chromatography (IC) (ICS2000, Dionex Inc., America). Dissolved organic carbon was analyzed using a total organic carbon analyzer (TOC-VCPH, Shimadzu Inc., Japan). Litterfall, needles and vegetation samples were washed by deionized water to remove the dust adsorbed on their surface, dried at 50 °C for 3 d, and then milled. Soil samples were air-dried at room temperature and milled. Total Hg concentrations in the solid samples above were analyzed using DMA80. The contents of C, N and S in organic matter of soil were measured by a dry combustion method (Vario EL cube, Elementar Inc., Germany). The information about QA/QC was given in Supplemental Information (SI). Hg2+ adsorption capacity of soil organic matter (Oe & Oa horizons) collected from each plot was measured in laboratory (see SI). The capacity of decomposed organic matter to bind Hg2+ is assessed by a distribution coefficient, D (mL g1) (Khwaja et al., 2006; Dong et al., 2011), defined as the ratio of the Hg adsorbed on the solid phase ([Hg2+]s, in ng g1) and the equilibrium Hg concentration in solution ([Hg2+]aq, in ng L1)
D ¼ ½Hg2þ s =½Hg2þ aq
ð1Þ
2.4. Data analysis Monitoring data of Hg, DOC and NO 3 concentrations and pH of throughfall and soil water were analyzed by repeated measures analysis of variance (RMANOVA) with the month (January 2010– December 2011) as the within-subjects factor and liming treatment as the between-subjects factor. One-way ANOVA was used for analyzing the differences of Hg content in major ground
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vegetation species, litterfall and pine needles, and C, N and S contents and Hg distribution coefficient (D) of decomposed organic matter. 3. Results 3.1. Hg in throughfall, soil, and soil water Although a large amount of Hg (121 lg m2, 20% soluble) was spread on the forest floor along with the gypsum application, the average Hg concentrations in the soil water of FGDG plots were decreased (by 17%, 8%, 37% and 43% for S0, S1, S2 and S3; only significantly in S2 and S3 by RMANOVA) over the two years (Fig. 1). Before treatment, there was no obvious difference of Hg content in soil among plots, with average values of 206 lg kg1 in soil Oe & Oa horizons, 166 lg kg1 in soil A horizon and 64 lg kg1 in soil B horizon. After treatment, the Hg content in soil Oe & Oa horizons of FGDG plots increased to 398 lg kg1 in November 2011, but in other horizons remained the same. No obvious difference in Hg concentrations in TF was detected between the two types of plots (Fig. 1), indicating that atmospheric deposition of Hg was similar (with the average value of 67.5 lg m2 yr1 similar with the total Hg input by the gypsum application (60.5 lg m2 yr1)). Compared with the Control plots, the gypsum application decreased the average concentrations of DOC and NO 3 in S1, S2, and S3 (significant in S1, S2 and S3 for DOC, and in S1 and S2 for NO 3 by RMANOVA) over the two years but increased the pH (significant in S1 by RMANOVA) (Fig. 1). 3.2. Hg in leaves and litterfall The average Hg contents in leaf tissues of ground vegetation in the FGDG plots were generally lower than in the Control plots (being 34%, 9%, 25% and 9% lower for W. japonica, D. fuscipes, D. pedata and M. sinensis; significant in W. japonica and D. pedata, Fig. 2). There were no significant differences in litterfall and needles between the two types of plots (Fig. 2). 3.3. Hg adsorption by organic matter The decomposed organic matter layers (Oe & Oa horizons) of the FGDG plots had significantly higher Hg distribution coefficients (D) than those of the Control plots, and the contents of N and S in the decomposed organic matter were significantly raised by the gypsum application (Fig. 3). However, no significant increase of C content was observed (Fig. 3). 4. Discussion 4.1. Inhibited soil Hg leaching to soil water by FGDG addition No significant increase of soil water Hg concentration in S0 was detected, suggesting that most of Hg with the gypsum application was retained in the organic layer. Previous studies found that organic matter showed strong capability of binding Hg because of the existence of numerous organic groups, especially the reduced organic sulfur sites (e.g., thiols (R–SH) and disulfide (R–SS–R)/disulfane (R–SSH)) (Laskowski et al., 1995; Hesterberg et al., 2001). The organic S concentration in the Oe & Oa horizons was 1.08 ± 0.07 g kg1 in the Control plots, and the value was significantly increased to 1.53 ± 0.18 g kg1 in the FGDG plots (Fig. 3), both of which were lower than the organic S concentration in the two studies mentioned before (1.65–2.68 g kg1; Laskowski et al., 1995; Hesterberg et al., 2001). Part of the organic S in organic matter was in reduced form (Khwaja et al., 2006).
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Fig. 1. The average Hg, DOC and NO 3 concentrations and pH of throughfall and soil water. ⁄, ⁄⁄ and ⁄⁄⁄ denote significant differences at the p < 0.05, p < 0.01 and p < 0.001 levels, respectively, according to a RMANOVA. All the statistical analysis results for the figure are listed in Table S2 (a) in SI. The abbreviations of the X axis represent throughfall (TF) and soil water (at different soil depths: S0, beneath organic matter; S1, 5 cm; S2, 15 cm; and S3, 30 cm).
Fig. 2. The average Hg contents in litterfall, needles of Masson pine and leaf tissues of four herbaceous plants in 2010 and 2011. ⁄ and ⁄⁄ denote significant differences at the p < 0.05 and p < 0.01 levels according to a one-way ANOVA. All the statistical analysis results for the figure are listed in Table S2 (b) in SI. WJ: Woodwardia japonica, DF: Dryopteris fuscipes, DP: Dicranopteris pedata, MS: Miscanthus sinensis.
Fig. 3. The contents of C, N, and S (g kg1) and Hg distribution coefficient (D) of organic matter (in Oe & Oa horizon). n = 9; ⁄ and ⁄⁄ denote significant differences between the two plot types at the p < 0.05 and p < 0.01 levels according to a oneway ANOVA; R2 represents the correlation between two elements.
The average dissolved organic carbon (DOC) concentrations in the soil water decreased in the FGDG plots (Fig. 1), probably due to the increased ionic strength caused by the dissolution of gypsum (Evans et al., 1988; Duan et al., 2008). Since DOC displayed a positive correlation with the Hg concentration in soil water (R2 = 0.38,
p < 0.01; Fig. S1 in SI; which agrees with other studies (e.g., Lindberg and Harriss, 1975; Driscoll et al., 1995; Ravichandran, 2004)), the lower DOC concentrations may be an important reason for the lower Hg concentrations in the soil water of the FGDG plots. Another reason may be the increased pH of soil water (Fig. 1) in the
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FGDG plots. The lower DOC concentration and higher pH also decreased the concentrations of other heavy metals (e.g. Cr, Cu, Pb, Ni and Cd) in soil water of the FGDG plots (Luo et al., 2012). The decrease in nitrate (NO 3 ) concentrations in the soil water of the FGDG plots (Fig. 1), which was as also found in previous studies (Belkacem and Nys, 1995), was probably due to the increase in N immobilization. 4.2. Inhibited soil Hg emission to the atmosphere by FGDG addition Although Hg emission flux from soil was not measured directly, lower Hg emission flux from soil was indicated by the lower Hg contents in leaf tissues of major herbaceous plants in the FGDG plots than in the Control plots (Fig. 2). Numerous studies have indicated that most of the Hg in leaves originates from the atmosphere as shown by the positive correlation between Hg concentrations in air and in leaf tissues (Ericksen et al., 2003; Ericksen and Gustin, 2004; Millhollen et al., 2006; Fay and Gustin, 2007). These studies using environmentally controlled growth chambers indicated that almost all of the Hg in leaf tissues was from air but not from soil. Leaf tissues of the ground vegetation can therefore reflect the Hg concentration in ambient air near forest soil. Since the space under forest canopy is relatively closed, the Hg concentration in air near forest soil would be greatly affected by soil emission. Soil emission may then affect the Hg content in ground vegetation which was very close to soil surface (averagely 40 cm in height). The Hg concentration in the air near forest floor (5 cm height) has been measured for two d (in November 4, 2010 and May 27, 2011). The results showed that the Hg concentration in the FGDG plots were significantly lower than that in the Control plots (Fig. S2 in SI), which was consistent with the observation results of leaf tissues. Because the direct measurements of Hg in the air were very limited, the Hg concentration in the leaf tissues of herbaceous plants would be more representative for Hg concentration in the air near forest floor. In contrast, the height of Masson pine in the forest is higher than 15 m, and the Hg content in litterfall and needles therefore could not be easily affected by soil emission from a relatively small plot, which resulting in no significant differences in Hg contents of litterfall and needles between the two types of plots. The elevated soil organic matter content, especially organic S content, in the surface soil of the FGDG plots (Fig. 3) was likely to be the main reason for the decrease of soil Hg emission. The reduction of oxidized mercury (Hg2+) to Hg0, induced by microorganisms (Mason et al., 1995), direct photolysis (Gustin et al., 2002), or humic substances (Alberts et al., 1974), mainly occurs in the surface soil (the organic matter layer). The increase of organic S content in the decomposed organic matter in the FGDG plots could retard soil Hg emission by enhancing the binding of
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Hg2+ by the soil (Hesterberg et al., 2001; Khwaja et al., 2006; Dong et al., 2011). In addition, Hg0 can be oxidized by thiol functional groups (–SH) via oxidative complexation (Zheng et al., 2012), directly decreasing Hg emission. Since both decomposed organic matter (solid phase) and dissolved organic matter (liquid phase) contain reduced organic S groups, the Hg concentration of equilibrium solutions in the adsorption experiment depends on the competitive bonding of Hg2+ by organic S groups in DOC versus in decomposed organic matter. The adsorption experiment indicated that the distribution coefficient (D) of Hg2+ had a significant positive correlation with the organic S content in decomposed organic matter, but a negative correlation with the DOC concentration in solution (Fig. 4). The organic S content in decomposed organic matter of FGDG plots was significantly raised, but no obvious change in DOC concentrations in solutions was observed (Fig. 4). Therefore, the enhancement of Hg2+-binding in the FGDG plots was mainly due to the increase of organic S content in the decomposed organic matter. The increase of organic S content was consistent with the increase of organic N content in the FGDG plots (Fig. 3), which means that the increased soil N immobilization caused by FGDG application resulted in increasing content of not only organic N, but also organic S in the soil. In organisms, most N and S are present in the form of proteins (Dijkshoorn and van Wijk, 1967), hence a positive correlation between N and S contents exists during the processes of organic matter decomposition (Tabatabai and Al-Khafaji, 1980), explaining the observed positive correlation between N and S contents in the soil (Fig. S3 in SI; R2 = 0.64, p = 0.027). 4.3. Hg Fluxes and pools Before the gypsum treatment, the Hg content in soil was very similar for all plots, and there was no obvious spatial difference in each plot. The gypsum applications introduced averagely 60.5 lg m2 yr1 Hg into the forest (with the doses of soluble Hg being about 11.5 lg m2 in 2009 and 12.7 lg m2 in 2011, similar as other researches (Hintelmann et al., 2002)), but the Hg fluxes in soil water were significantly lower in the FGDG plots compared to the Control plots (Fig. 5), suggesting that most of the Hg in the gypsum was retained in the organic matter layer. Compared with the Control plots, the amount of Hg stored in Oe & Oa horizons in the FGDG plots increased by 71.0 lg m2 yr1 within the two years (Fig. 5). The decreased Hg flux in the S0 soil water contributed 2.3 lg m2 yr1 to the Hg pool in the Oe & Oa horizons, and the other 8.2 lg m2 yr1 (71.0 60.5 2.3 = 8.2) of the pool may be attributable to the decrease in Hg emissions from the soil in the FGDG plots (since the Hg input by litterfall and throughfall did
Fig. 4. Distribution coefficient (D) versus organic S in the solid phase (left) and DOC in the liquid phase (right). n = 9; ⁄ and ⁄⁄ denote significant differences between the two plot types at the p < 0.05 and p < 0.01 levels according to a one-way ANOVA.
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Fig. 5. Hg fluxes (two year average) and pools (two years after FGDG applications) in the forest. Fluxes (lg m2 yr1) are represented by arrows calculated by multiplying Hg concentrations and mass of water or litterfall, and pools (lg m2) are shown in the white boxes calculated by multiplying Hg concentrations and biomass of masson pine needle or bulk densities of organic matter or soil. ⁄ and ⁄⁄ denote significant differences between the two plot types at the p < 0.05 and p < 0.01 levels according to a one-way ANOVA.
not change). Compared to the reported Hg emission flux from soil in the Control plots (73.6 lg m2 yr1, about 82% of the total input by litterfall and throughfall) (Wang et al., 2009), the emission flux from the soil was reduced by approximately 11% by the gypsum application in this study. 4.4. Environmental implications For the acid forest soils primarily in southern China, high deposition of Ca2+ and SO2 4 during the past three decades may lower Hg release from soil (both emission and leaching). Because some of the natural Hg sources are re-emissions of deposited Hg from anthropogenic emissions (Lindberg et al., 2007), the effect of atmospheric deposition may lower the net anthropogenic Hg emissions from China which is involved in global Hg transport. However, this high deposition, correlated with serious pollution of PM and acid deposition, is likely to decrease in the near future. To decrease the emission of acid pollutants mainly for improving air quality, the Chinese government has set compulsory SO2 and PM emission limits (Zhao et al., 2011). Accompanied by PM emission control applied to the cement industry and other anthropogenic sources, Ca2+ deposition will rapidly decrease in the future in China, and thus delay the recovery from acidification by emission abatement of SO2 and NOx (Zhao et al., 2011). At the same time, the risk of increasing Hg emission from soil can also be expected.
input to the forest soil through FGDG, the gypsum treatment reduced the Hg concentration in soil water, as well as the soil Hg emission (shown by the lower Hg contents in leaf tissues of ground vegetation). Most of the Hg in the gypsum was retained in the soil Oe & Oa horizons. The decreased DOC concentration in soil water and the elevated organic sulfur content in surface soil after the gypsum treatment were speculated to be the reasons for the decrease in Hg release from the soil. Soil Hg emission is likely to increase in the future since the atmospheric deposition of Ca2+ and SO2 is expected to decrease with the stringent control of 4 PM and SO2 in China. Acknowledgements The research is financially supported by the National Natural Science Foundation of China (Nos. 20877047, 21221004, and 21377064).
Appendix A. Supplementary material Supplementary data associated with this article can be found, in the online version, at http://dx.doi.org/10.1016/j.chemosphere. 2015.03.081. References
5. Conclusions This study investigated the effect of high atmospheric deposition of Ca2+ and SO2 4 on the Hg release from forest soil by applying the gypsum (FGDG) on the soil. Although a large amount of Hg was
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Inhibition of mercury release from forest soil by high atmospheric deposition of Ca2+ and SO2 4 Yao Luo a, Lei Duan a,b,⇑, Guangyi Xu c, Jiming Hao a,b a
State Key Laboratory of Environmental Simulation and Pollution Control, School of Environment, Tsinghua University, Beijing 100084, China Collaborative Innovation Centre for Regional Environmental Quality, Tsinghua University, Beijing 100084, China c Shenzhen Academy of Environment Sciences, Shenzhen 518001, China b
h i g h l i g h t s 2+
Hg release from forest soil was inhibited by high deposition of Ca
and SO2 4 .
The elevated organic S enhanced the capacity of surface soil to bind Hg. Soil Hg emission may increase in the future due to stringent control of PM and SO2.
a r t i c l e
i n f o
Article history: Received 1 July 2014 Received in revised form 14 March 2015 Accepted 23 March 2015 Available online 15 May 2015 Handling Editor: X. Cao Keywords: Flue gas desulfurization gypsum (FGDG) Mercury Natural source Forest soil Organic sulfur Atmospheric deposition
a b s t r a c t As one of the most important natural mercury (Hg) sources, soil release (emission to the atmosphere or leaching to soil water) depends on various factors, some of which can be affected by atmospheric deposition. We studied the effect of flue gas desulfurization gypsum (FGDG) addition on soil Hg release in a Masson pine (Pinus massoniana) forest in southwestern China. FGDG addition simulated atmospheric deposition of Ca2+, SO2 4 and Hg, which are commonly high in China. Results showed that Hg concentration in soil water decreased with the gypsum treatment, suggesting that the mobility of Hg in mineral soil was reduced. Moreover, the application of gypsum also seems to have decreased Hg emission from the soil, shown by the lower Hg contents in leaf tissues of ground vegetation in the treated plots than in the reference. Both Hg mobility in the soil and Hg emission to the atmosphere were decreased despite the additional Hg input from FGDG. The decreased DOC concentration in soil water and the elevated organic sulfur content in the soil Oe & Oa horizons were speculated to result in an enhanced capacity of surface soil to bind Hg, and thus to reduce Hg release from the soil. However, with the increasingly stringent control of particulate matter (PM) and sulfur dioxide (SO2) emissions in China, the deposition of Ca2+ and SO2 4 is expected to decrease, and their ability to inhibit soil Hg release is likely to decline in the future. Ó 2015 Elsevier Ltd. All rights reserved.
1. Introduction Mercury pollution is a global environmental concern due to its harm to human beings (Jensen and Jernelöv, 1969; Mergler et al., 2007). In terrestrial ecosystem, forest soil is an important natural source of Hg. Not only can Hg be leached out from forest soil through soil water (Hintelmann et al., 2002), but also large amounts of Hg emit to the atmosphere (Lindberg et al., 1979; Xiao et al., 1991). The soil properties, such as pH value, Hg content ⇑ Corresponding author at: State Key laboratory of Environmental Simulation and Pollution Control, School of Environment, Tsinghua University, Beijing 100084, China. Tel.: +86 (10)62783758; fax: +86 (10)62773957. E-mail address: [email protected] (L. Duan). http://dx.doi.org/10.1016/j.chemosphere.2015.03.081 0045-6535/Ó 2015 Elsevier Ltd. All rights reserved.
and humic matter content, may influence Hg release significantly (Gillis and Miller, 2000; Mauclair et al., 2008). As soil properties can be greatly affected by atmospheric deposition (e.g., soil acidification), it is important to clarify the effects of atmospheric deposition on forest soil Hg release. China is currently the world’s top emitter of anthropogenic Hg due to enormous coal combustion, nonferrous metals production and so on (Pirrone et al., 2010; Wang et al., 2010; Wu et al., 2012), which results in high Hg deposition in China. The Hg deposition in remote areas and urban areas in China were 1–2 times and 1–2 magnitude respectively higher than those in North America and Europe (Fu et al., 2012). Recent estimation using CMAQ-Hg model showed that dry deposition of Hg in the east of China in 2005 was generally more than 40 lg m2 yr1, and wet deposition
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more than 20 lg m2 yr1 (Lin et al., 2010), much higher than the total Hg deposition in northeastern North America (21.3 lg m2 yr1; Miller et al., 2005) and in Japan (20.8 lg m2 yr1; Sakata and Marumoto, 2005). Large amounts of Hg in soil deposited from atmosphere would be harmful to human beings or reemitted from soil. Due to serious pollution of particulate matter (PM) and sulfur dioxide (SO2), atmospheric deposition in China has also high con-
1971 to 2000 (Meteorological Station of Chongqing), with an annual average precipitation pH of 4.0–4.2 (Larssen et al., 2011). Atmospheric deposition of sulfur (8–10 keq ha1 yr1), nitrogen (about 3 keq ha1 yr1, with two-thirds NH+4 and one-third NO 3) and Ca2+ (about 6 keq ha1 yr1) is high and the soil is a Haplic Acrisol which is acidic (pH 3.6–4.2) (Larssen et al., 2011).
2+ tents of Ca2+ and SO2 and SO2 4 . The deposition fluxes of Ca 4 at some monitored sites in southern and southwestern China (Ca2+:
2.2. FGDG component analysis and application
1 0.8–5.7 keq ha1 yr1, SO2 yr1; Larssen et al., 4 : 1.5–10.5 keq ha 2011) are much higher than those in North America and Europe during the peak period of acid deposition. This high deposition may change soil chemical properties and then influence the soil Hg release. With the increasingly stringent control of PM and SO2
The gypsum (FGDG) used in this study was obtained from a coal-fired power plant in Chongqing city. Total Hg content of the gypsum was analyzed by a direct Hg analyzer (DMA80, Milestone Inc., Italy), and Soluble Hg content was measured by sulfuric acid leaching (see SI). The contents of other elements (e.g. Ca and S) were measured by ICP-AES (Thermo Jarrell Ash, USA) after being decomposed by concentrated nitric acid. Carbonate (CaCO3) content was quantified using a titration method. The gypsum (FGDG) used in this study was about 88% pure (CaSO42H2O), with 8% of CaCO3 due to the use of excessive desulfurizing agent (limestone) in the FGD process. The amount of CaCO3 addition was approximately 1.00 keq ha1 yr1, which could neutralize about 10% of sulfur deposition. The total Hg concentration was 910 ± 10 lg kg1, and around 20% was soluble Hg. The amount of soluble Hg did not vary with water pH according to the leaching experiments (Table S1 in SI), indicating that the insoluble Hg in FGDG was quite stable against acid and might be mercuric sulfide (HgS) (Rallo et al., 2010). In October 2009, six 10 10 m2 plots were established in a homogeneous forest. The distance between the nearest plots was about 10 m. The plots were randomly divided into two groups, each with three replicates. One group was treated by the gypsum (FGDG) and the other was used as a control, denoted as FGDG and Control, respectively. The gypsum was sprinkled onto the forest floor manually at a dose of 63.4 g m2 in December 2009 and then at a dose of 70.0 g m2 in May 2011 for each FGDG plot. The dose was chosen to approximately double the annual deposition of Ca2+ in this area. Buffer zone with 1 m width around each FGDG plot was also treated. The visible layer of the gypsum disappeared soon after rain and did not influence the air/surface exchange.
emissions in China, however, the deposition of Ca2+ and SO2 4 and its influence on the Hg release are expected to decrease in the future (Zhao et al., 2011). In order to study the effects of the high deposition of Ca2+ and SO2 4 on soil Hg release, we applied flue gas desulfurization gypsum (FGDG), which is generated in the process of flue gas desulfurization (FGD) and mainly consists of CaSO4 and a little residual CaCO3, on forest soil to simulate the additional atmospheric deposition of Ca2+, SO2 and Hg. Gypsum (CaSO42H2O) was used 4 solution because more than two thirds instead of Ca2+ and SO2 4 of the deposition of Ca2+ and SO2 4 in China is through dry deposiin atmotion (Larssen et al., 2011), and most of Ca2+ and SO2 4 spheric particles exists as gypsum (Takahashi et al., 2008; Quan et al., 2008; Ma et al., 2013). The gypsum also contains some Hg, because Hg in flue gas can be transferred into the gypsum during the FGD process (Kairies et al., 2006; Wang et al., 2010). After the gypsum application, the decrease in Hg concentration in soil water was expected to be found, because the gypsum could decrease DOC concentration in soil water which showed strong positive correlation to Hg concentration in water (Evans et al., 1988; Akerblom et al., 2008). However, the change of soil Hg0 emission was uncertain. On one side, Hg0 emission may be increased because of the additional Hg inputs (Coolbaugh et al., 2002); on the other side, gypsum could elevate organic nitrogen content in soil (Belkacem and Nys, 1995) which could enhance the capability of soil to bind Hg and decrease Hg0 emission. Hg leaching in soil water and Hg content in soil and vegetation in the forest were investigated during 2010–2011. Although Hg emission flux from soil was not measured directly, the change of flux was reflected by the change of Hg content in leaf tissues of ground vegetation, because numerous studies have indicated that most of the Hg in leaves originates from the atmosphere (Ericksen et al., 2003; Ericksen and Gustin, 2004; Millhollen et al., 2006; Fay and Gustin, 2007). Here we hypothesize: (1) Hg concentration in soil water will be decreased by the gypsum application; and (2) the soil Hg0 emission will be influenced by the changes of soil chemical properties, which can be reflected in leaves tissues. 2. Materials and methods 2.1. Site description The field manipulation experiment was carried out in a Masson pine (Pinus massoniana) forest at Tieshanping (29°380 N, 106°410 E), about 25 km northeast of Chongqing City in southwestern China. Masson pine forests, either natural or planted, are widely distributed in the humid subtropical areas of southern China (Wu, 1980). The annual mean temperature and precipitation at the experimental site were 18.2 °C and 1105 mm respectively from
2.3. Sample collection and chemical analysis In October 2009, four throughfall (TF) collectors were placed in the corners of each plot, and two series of lysimeters were installed close to the center of each plot. The throughfall collector consisted of a funnel with a certain sectional area to collect throughfall and a bottle to store water. The material of the two sections was PET (polyethylene terephthalate), which was found to be hard to absorb mercury (Parker and Bloom, 2005). Each series of lysimeters contained a zero-tension plate lysimeters (0.30 0.30 m2) installed beneath the organic layer (S0), and three rhizons (which are made of inert polymer with very low metal absorption and suitable for trace metal research; Knight et al., 1998) installed at 5, 15, and 30 cm depth in the mineral soil (layers S1, S2, and S3). One series was used for Hg measurement, the sample bottles of which were pre-added with 4 mL of 6 mol L1-ultrapure HCl to inhibit the transformation and absorption of Hg in the bottles. The other series was used for measurements of pH value, ion concentrations and dissolved organic carbon (DOC). Since the beginning of 2010, throughfall and soil water from each layer were gathered weekly and refrigerated immediately. The samples of every four weeks were pooled into a bulk sample (monthly sample) for chemical analysis.
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Litterfall in each plot was gathered monthly from four 1 1 m2 square collectors, each close to a throughfall collector. Herbaceous ground vegetation was examined in five permanently marked 1 1 m2 quadrates in each plot. The five quadrates were distributed in such a way that each quadrat contained at least one of the major species in the forest (Woodwardia japonica, Dryopteris fuscipes, Dicranopteris pedata or Miscanthus sinensis), and all major species were included in the five quadrates. Only current-year leaves of those four species, which is generally 30–60 cm in height, were collected from each quadrate for chemical analysis. For fern (W. japonica, D. fuscipes and D. pedata), three pinnately compound leaves were completely cut by a steel scissor; for grass (M. sinensis), at least five leaves were cut. To measure needle Hg content, three branches were cut randomly from the crown top of each of the two highest Masson pines in each plot (15–25 m), and 1000 current-year needles were gathered from the branches. The leaves sampling was conducted in early November (at the end of the growing season) of each year. Soil was collected in October 2009 (before treatment) and October 2011 (about two years after treatment) from six layers of each plot, i.e., Oi horizon (undecomposed litter layer), Oe & Oa horizons (moderately decomposed organic matter layer & humus layer), A horizon (mineral soil 0–2 cm) and B1–B3 horizons (mineral soil 2–5 cm, 5–15 cm, and 15–30 cm, respectively). In each plot, three soil profiles were sampled, and three samples from the same layer were mixed to one sample. Organic soil was picked from forest floor using clean latex gloves and stainless spade, while mineral soil was collected using soil auger. All soil and plant samples were immediately bagged in plastic bags and stored in an insulation box with ice bags until transport to laboratory. After pH was analyzed, water samples were filtered through a 0.45-lm glass fiber filter. Total dissolved Hg concentration was measured by cold vapor atomic fluorescence spectrometry (CVAFS) (Tekran model 2600 system, Tekran Inc., Canada), following the modified U.S. EPA Method 1631 guideline (Holmes and þ Lean, 2006). Concentrations of other ions (e.g. NO 3 and NH4 ) were measured by ion chromatography (IC) (ICS2000, Dionex Inc., America). Dissolved organic carbon was analyzed using a total organic carbon analyzer (TOC-VCPH, Shimadzu Inc., Japan). Litterfall, needles and vegetation samples were washed by deionized water to remove the dust adsorbed on their surface, dried at 50 °C for 3 d, and then milled. Soil samples were air-dried at room temperature and milled. Total Hg concentrations in the solid samples above were analyzed using DMA80. The contents of C, N and S in organic matter of soil were measured by a dry combustion method (Vario EL cube, Elementar Inc., Germany). The information about QA/QC was given in Supplemental Information (SI). Hg2+ adsorption capacity of soil organic matter (Oe & Oa horizons) collected from each plot was measured in laboratory (see SI). The capacity of decomposed organic matter to bind Hg2+ is assessed by a distribution coefficient, D (mL g1) (Khwaja et al., 2006; Dong et al., 2011), defined as the ratio of the Hg adsorbed on the solid phase ([Hg2+]s, in ng g1) and the equilibrium Hg concentration in solution ([Hg2+]aq, in ng L1)
D ¼ ½Hg2þ s =½Hg2þ aq
ð1Þ
2.4. Data analysis Monitoring data of Hg, DOC and NO 3 concentrations and pH of throughfall and soil water were analyzed by repeated measures analysis of variance (RMANOVA) with the month (January 2010– December 2011) as the within-subjects factor and liming treatment as the between-subjects factor. One-way ANOVA was used for analyzing the differences of Hg content in major ground
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vegetation species, litterfall and pine needles, and C, N and S contents and Hg distribution coefficient (D) of decomposed organic matter. 3. Results 3.1. Hg in throughfall, soil, and soil water Although a large amount of Hg (121 lg m2, 20% soluble) was spread on the forest floor along with the gypsum application, the average Hg concentrations in the soil water of FGDG plots were decreased (by 17%, 8%, 37% and 43% for S0, S1, S2 and S3; only significantly in S2 and S3 by RMANOVA) over the two years (Fig. 1). Before treatment, there was no obvious difference of Hg content in soil among plots, with average values of 206 lg kg1 in soil Oe & Oa horizons, 166 lg kg1 in soil A horizon and 64 lg kg1 in soil B horizon. After treatment, the Hg content in soil Oe & Oa horizons of FGDG plots increased to 398 lg kg1 in November 2011, but in other horizons remained the same. No obvious difference in Hg concentrations in TF was detected between the two types of plots (Fig. 1), indicating that atmospheric deposition of Hg was similar (with the average value of 67.5 lg m2 yr1 similar with the total Hg input by the gypsum application (60.5 lg m2 yr1)). Compared with the Control plots, the gypsum application decreased the average concentrations of DOC and NO 3 in S1, S2, and S3 (significant in S1, S2 and S3 for DOC, and in S1 and S2 for NO 3 by RMANOVA) over the two years but increased the pH (significant in S1 by RMANOVA) (Fig. 1). 3.2. Hg in leaves and litterfall The average Hg contents in leaf tissues of ground vegetation in the FGDG plots were generally lower than in the Control plots (being 34%, 9%, 25% and 9% lower for W. japonica, D. fuscipes, D. pedata and M. sinensis; significant in W. japonica and D. pedata, Fig. 2). There were no significant differences in litterfall and needles between the two types of plots (Fig. 2). 3.3. Hg adsorption by organic matter The decomposed organic matter layers (Oe & Oa horizons) of the FGDG plots had significantly higher Hg distribution coefficients (D) than those of the Control plots, and the contents of N and S in the decomposed organic matter were significantly raised by the gypsum application (Fig. 3). However, no significant increase of C content was observed (Fig. 3). 4. Discussion 4.1. Inhibited soil Hg leaching to soil water by FGDG addition No significant increase of soil water Hg concentration in S0 was detected, suggesting that most of Hg with the gypsum application was retained in the organic layer. Previous studies found that organic matter showed strong capability of binding Hg because of the existence of numerous organic groups, especially the reduced organic sulfur sites (e.g., thiols (R–SH) and disulfide (R–SS–R)/disulfane (R–SSH)) (Laskowski et al., 1995; Hesterberg et al., 2001). The organic S concentration in the Oe & Oa horizons was 1.08 ± 0.07 g kg1 in the Control plots, and the value was significantly increased to 1.53 ± 0.18 g kg1 in the FGDG plots (Fig. 3), both of which were lower than the organic S concentration in the two studies mentioned before (1.65–2.68 g kg1; Laskowski et al., 1995; Hesterberg et al., 2001). Part of the organic S in organic matter was in reduced form (Khwaja et al., 2006).
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Fig. 1. The average Hg, DOC and NO 3 concentrations and pH of throughfall and soil water. ⁄, ⁄⁄ and ⁄⁄⁄ denote significant differences at the p < 0.05, p < 0.01 and p < 0.001 levels, respectively, according to a RMANOVA. All the statistical analysis results for the figure are listed in Table S2 (a) in SI. The abbreviations of the X axis represent throughfall (TF) and soil water (at different soil depths: S0, beneath organic matter; S1, 5 cm; S2, 15 cm; and S3, 30 cm).
Fig. 2. The average Hg contents in litterfall, needles of Masson pine and leaf tissues of four herbaceous plants in 2010 and 2011. ⁄ and ⁄⁄ denote significant differences at the p < 0.05 and p < 0.01 levels according to a one-way ANOVA. All the statistical analysis results for the figure are listed in Table S2 (b) in SI. WJ: Woodwardia japonica, DF: Dryopteris fuscipes, DP: Dicranopteris pedata, MS: Miscanthus sinensis.
Fig. 3. The contents of C, N, and S (g kg1) and Hg distribution coefficient (D) of organic matter (in Oe & Oa horizon). n = 9; ⁄ and ⁄⁄ denote significant differences between the two plot types at the p < 0.05 and p < 0.01 levels according to a oneway ANOVA; R2 represents the correlation between two elements.
The average dissolved organic carbon (DOC) concentrations in the soil water decreased in the FGDG plots (Fig. 1), probably due to the increased ionic strength caused by the dissolution of gypsum (Evans et al., 1988; Duan et al., 2008). Since DOC displayed a positive correlation with the Hg concentration in soil water (R2 = 0.38,
p < 0.01; Fig. S1 in SI; which agrees with other studies (e.g., Lindberg and Harriss, 1975; Driscoll et al., 1995; Ravichandran, 2004)), the lower DOC concentrations may be an important reason for the lower Hg concentrations in the soil water of the FGDG plots. Another reason may be the increased pH of soil water (Fig. 1) in the
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FGDG plots. The lower DOC concentration and higher pH also decreased the concentrations of other heavy metals (e.g. Cr, Cu, Pb, Ni and Cd) in soil water of the FGDG plots (Luo et al., 2012). The decrease in nitrate (NO 3 ) concentrations in the soil water of the FGDG plots (Fig. 1), which was as also found in previous studies (Belkacem and Nys, 1995), was probably due to the increase in N immobilization. 4.2. Inhibited soil Hg emission to the atmosphere by FGDG addition Although Hg emission flux from soil was not measured directly, lower Hg emission flux from soil was indicated by the lower Hg contents in leaf tissues of major herbaceous plants in the FGDG plots than in the Control plots (Fig. 2). Numerous studies have indicated that most of the Hg in leaves originates from the atmosphere as shown by the positive correlation between Hg concentrations in air and in leaf tissues (Ericksen et al., 2003; Ericksen and Gustin, 2004; Millhollen et al., 2006; Fay and Gustin, 2007). These studies using environmentally controlled growth chambers indicated that almost all of the Hg in leaf tissues was from air but not from soil. Leaf tissues of the ground vegetation can therefore reflect the Hg concentration in ambient air near forest soil. Since the space under forest canopy is relatively closed, the Hg concentration in air near forest soil would be greatly affected by soil emission. Soil emission may then affect the Hg content in ground vegetation which was very close to soil surface (averagely 40 cm in height). The Hg concentration in the air near forest floor (5 cm height) has been measured for two d (in November 4, 2010 and May 27, 2011). The results showed that the Hg concentration in the FGDG plots were significantly lower than that in the Control plots (Fig. S2 in SI), which was consistent with the observation results of leaf tissues. Because the direct measurements of Hg in the air were very limited, the Hg concentration in the leaf tissues of herbaceous plants would be more representative for Hg concentration in the air near forest floor. In contrast, the height of Masson pine in the forest is higher than 15 m, and the Hg content in litterfall and needles therefore could not be easily affected by soil emission from a relatively small plot, which resulting in no significant differences in Hg contents of litterfall and needles between the two types of plots. The elevated soil organic matter content, especially organic S content, in the surface soil of the FGDG plots (Fig. 3) was likely to be the main reason for the decrease of soil Hg emission. The reduction of oxidized mercury (Hg2+) to Hg0, induced by microorganisms (Mason et al., 1995), direct photolysis (Gustin et al., 2002), or humic substances (Alberts et al., 1974), mainly occurs in the surface soil (the organic matter layer). The increase of organic S content in the decomposed organic matter in the FGDG plots could retard soil Hg emission by enhancing the binding of
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Hg2+ by the soil (Hesterberg et al., 2001; Khwaja et al., 2006; Dong et al., 2011). In addition, Hg0 can be oxidized by thiol functional groups (–SH) via oxidative complexation (Zheng et al., 2012), directly decreasing Hg emission. Since both decomposed organic matter (solid phase) and dissolved organic matter (liquid phase) contain reduced organic S groups, the Hg concentration of equilibrium solutions in the adsorption experiment depends on the competitive bonding of Hg2+ by organic S groups in DOC versus in decomposed organic matter. The adsorption experiment indicated that the distribution coefficient (D) of Hg2+ had a significant positive correlation with the organic S content in decomposed organic matter, but a negative correlation with the DOC concentration in solution (Fig. 4). The organic S content in decomposed organic matter of FGDG plots was significantly raised, but no obvious change in DOC concentrations in solutions was observed (Fig. 4). Therefore, the enhancement of Hg2+-binding in the FGDG plots was mainly due to the increase of organic S content in the decomposed organic matter. The increase of organic S content was consistent with the increase of organic N content in the FGDG plots (Fig. 3), which means that the increased soil N immobilization caused by FGDG application resulted in increasing content of not only organic N, but also organic S in the soil. In organisms, most N and S are present in the form of proteins (Dijkshoorn and van Wijk, 1967), hence a positive correlation between N and S contents exists during the processes of organic matter decomposition (Tabatabai and Al-Khafaji, 1980), explaining the observed positive correlation between N and S contents in the soil (Fig. S3 in SI; R2 = 0.64, p = 0.027). 4.3. Hg Fluxes and pools Before the gypsum treatment, the Hg content in soil was very similar for all plots, and there was no obvious spatial difference in each plot. The gypsum applications introduced averagely 60.5 lg m2 yr1 Hg into the forest (with the doses of soluble Hg being about 11.5 lg m2 in 2009 and 12.7 lg m2 in 2011, similar as other researches (Hintelmann et al., 2002)), but the Hg fluxes in soil water were significantly lower in the FGDG plots compared to the Control plots (Fig. 5), suggesting that most of the Hg in the gypsum was retained in the organic matter layer. Compared with the Control plots, the amount of Hg stored in Oe & Oa horizons in the FGDG plots increased by 71.0 lg m2 yr1 within the two years (Fig. 5). The decreased Hg flux in the S0 soil water contributed 2.3 lg m2 yr1 to the Hg pool in the Oe & Oa horizons, and the other 8.2 lg m2 yr1 (71.0 60.5 2.3 = 8.2) of the pool may be attributable to the decrease in Hg emissions from the soil in the FGDG plots (since the Hg input by litterfall and throughfall did
Fig. 4. Distribution coefficient (D) versus organic S in the solid phase (left) and DOC in the liquid phase (right). n = 9; ⁄ and ⁄⁄ denote significant differences between the two plot types at the p < 0.05 and p < 0.01 levels according to a one-way ANOVA.
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Fig. 5. Hg fluxes (two year average) and pools (two years after FGDG applications) in the forest. Fluxes (lg m2 yr1) are represented by arrows calculated by multiplying Hg concentrations and mass of water or litterfall, and pools (lg m2) are shown in the white boxes calculated by multiplying Hg concentrations and biomass of masson pine needle or bulk densities of organic matter or soil. ⁄ and ⁄⁄ denote significant differences between the two plot types at the p < 0.05 and p < 0.01 levels according to a one-way ANOVA.
not change). Compared to the reported Hg emission flux from soil in the Control plots (73.6 lg m2 yr1, about 82% of the total input by litterfall and throughfall) (Wang et al., 2009), the emission flux from the soil was reduced by approximately 11% by the gypsum application in this study. 4.4. Environmental implications For the acid forest soils primarily in southern China, high deposition of Ca2+ and SO2 4 during the past three decades may lower Hg release from soil (both emission and leaching). Because some of the natural Hg sources are re-emissions of deposited Hg from anthropogenic emissions (Lindberg et al., 2007), the effect of atmospheric deposition may lower the net anthropogenic Hg emissions from China which is involved in global Hg transport. However, this high deposition, correlated with serious pollution of PM and acid deposition, is likely to decrease in the near future. To decrease the emission of acid pollutants mainly for improving air quality, the Chinese government has set compulsory SO2 and PM emission limits (Zhao et al., 2011). Accompanied by PM emission control applied to the cement industry and other anthropogenic sources, Ca2+ deposition will rapidly decrease in the future in China, and thus delay the recovery from acidification by emission abatement of SO2 and NOx (Zhao et al., 2011). At the same time, the risk of increasing Hg emission from soil can also be expected.
input to the forest soil through FGDG, the gypsum treatment reduced the Hg concentration in soil water, as well as the soil Hg emission (shown by the lower Hg contents in leaf tissues of ground vegetation). Most of the Hg in the gypsum was retained in the soil Oe & Oa horizons. The decreased DOC concentration in soil water and the elevated organic sulfur content in surface soil after the gypsum treatment were speculated to be the reasons for the decrease in Hg release from the soil. Soil Hg emission is likely to increase in the future since the atmospheric deposition of Ca2+ and SO2 is expected to decrease with the stringent control of 4 PM and SO2 in China. Acknowledgements The research is financially supported by the National Natural Science Foundation of China (Nos. 20877047, 21221004, and 21377064).
Appendix A. Supplementary material Supplementary data associated with this article can be found, in the online version, at http://dx.doi.org/10.1016/j.chemosphere. 2015.03.081. References
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