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Published online 20 February 2008
Published in J Environ Qual 37:680-688 (2008)
DOI: 10.2134/jeq2007.0221
© 2008 American Society of Agronomy, Crop Science Society of America, and Soil Science Society of America
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TECHNICAL REPORTS

Heavy Metals in the Environment

Adsorption–Desorption Characteristics of Mercury in Paddy Soils of China

Y. D. Jinga,b, Z. L. Hea,c,* and X. E. Yanga

a MOE Key Lab. of Environ., Remediation and Ecosystem Health, College of Natural Resource and Environment Sciences, Zhejiang Univ., Hangzhou 310029, China
b Dep. of Resources and Planning, Qufu Normal Univ., Jining 273165, China
c Univ. of Florida, Inst. of Food and Agricultural Sciences, Indian River Research and Education Center, Fort Pierce, FL 34945, USA

* Corresponding author (zhe{at}ufl.edu).

Received for publication May 2, 2007.

   ABSTRACT
TOP
 NOTES
 ABSTRACT
INTRODUCTION
 Materials and Methods
 Results
 Discussion
 Conclusions
 REFERENCES
 
Mercury (Hg) has received considerable attention because of its association with various human health problems. Adsorption–desorption behavior of Hg at contaminated levels in two paddy soils was investigated. The two representative soils for rice production in China, locally referred to as a yellowish red soil (YRS) and silty loam soil (SLS) and classified as Gleyi-Stagnic Anthrosols in FAO/UNESCO nomenclature, were respectively collected from Jiaxin County and Xiasha District of Hangzhou City, Zhejiang Province. The YRS adsorbed more Hg2+ than the SLS. The characteristics of Hg adsorption could be described by the simple Langmuir adsorption equation (r2 = 0.999 and 0.999, P < 0.01, respectively, for the SLS and YRS). The maximum adsorption values (Xm) that were obtained from the simple Langmuir model were 111 and 213 mg Hg2+ kg–1 soil, respectively, for the SLS and YRS. Adsorption of Hg2+ decreased soil pH by 0.75 unit for the SLS soil and 0.91 unit for the YRS soil at the highest loading. The distribution coefficient (kd) of Hg in the soil decreased exponentially with increasing Hg2+ loading. After five successive desorptions with 0.01 mol L–1 KCl solution (pH 5.4), 0 to 24.4% of the total adsorbed Hg2+ in the SLS soil was desorbed and the corresponding value of the YRS soil was 0 to 14.4%, indicating that the SLS soil had a lower affinity for Hg2+ than the YRS soil at the same Hg2+ loading. Different mechanisms are likely involved in Hg2+ adsorption–desorption at different levels of Hg2+ loading and between the two soils.

Abbreviations: CEC, cation exchange capacity • SLS, silty loam soil • YRS, yellowish red soil


   INTRODUCTION
TOP
 NOTES
 ABSTRACT
 INTRODUCTION
Materials and Methods
 Results
 Discussion
 Conclusions
 REFERENCES
 
MERCURY is a worldwide environmental hazard. This is due, in part, to Hg dispersion on a global scale, its extreme toxicity at low levels, and its tendency to bioaccumulate in organisms and biomagnify in food chains (Mastrine et al., 1999). Since the outbreak of Minamata disease in the 1950s it has become known that Hg represents both a potential and actual threat to human health (WHO, 1976). It is listed as one of the 126 priority contaminants by the United States Environmental Protection Agency (USEPA) as a consequence of its accumulative and persistent character in the environment and biota.

Human use of Hg has a history of more than 2000 yr, with a massive increase occurring during the industrial revolution. In the last 100 yr, global atmospheric emissions of Hg have increased by a factor of 3 (Hanish, 1998). Recent estimates indicate that approximately 95% of Hg emitted from anthropogenic sources resides in terrestrial soils (USEPA, 1997). The major sources of mercury poisoning are its mining and recovery and its usage in a variety of products, industrial waste, and accumulation in the human body via the food chain.

Mercury and its compounds are hazardous for humans, plants, and animals. Mercury metal, its vapor, and most of its organic and inorganic compounds are protoplasmic poisons. The tolerance limit of inorganic Hg in aqueous solution is 1 mg L–1. Contamination of the aquatic food chain of the Madeira River tributaries has been demonstrated by Maurice-Bourgoin et al. (2000), who report that 86% of the piscivorous fishes collected in the Beni River exhibit Hg concentrations at values exceeding almost four times the WHO safety limit (WHO, 1976).

Total mercury in soil is generally low. However, there were reports which recorded very high Hg level. Mercury concentration in the active Hg mining areas in Guizhou, China ranged from 90 to 150 mg kg–1 and in Alaska the reported total Hg levels were as high as 3200 mg kg–1 (NIMD, 2000). Gray et al. (2003) documented concentrations of 4100 mg kg–1 in surficial sediments surrounding the Sitio Honda Bay jetty and 43 to 660 mg kg–1 in mine waste calcines, respectively. Mercury concentration of soil samples collected in the surroundings of a chlor-alkali plant in the Netherlands varied from 4.3 to 1150 mg kg–1. High Hg concentrations (>1.5 mg kg–1) have been reported in fishes from Florida Bay near the Everglades, in South Florida (Evans and Crumley, 2005). Efforts have been previously devoted to the study of low Hg level contamination (WHO, 1976; Yin et al., 1997; Miretzky et al., 2005), but data on adsorption–desorption behavior of Hg in highly contaminated soils or environments are needed for management of high-Hg environments.

Mercury adsorbed onto soil is subjected to a wide array of chemical and biological transformation processes such as Hg (0) oxidation and Hg (II) reduction or methylation, depending on soil pH, temperature, and soil organic matter content (Weber, 1993). An overview of Hg binding in soils is given by Schuster (1991). It is known that Hg mobilization in soils through formation of inorganic soluble Hg compounds such as HgCl2 and Hg(OH)2 are of minor importance in the presence of organic matter, as Hg is known to be effectively bound to soil humic substances (Weber, 1988). The formation of organic Hg (II) complexes is known to be the dominating process, which is due largely to the affinity of Hg (II) and its inorganic compounds to sulfur-containing functional groups (Schuster, 1991; Yin et al., 1997). Although a large number of laboratory and field studies have examined the dynamic equilibrium of solid and liquid phases of Hg in the environment (Scholtz et al., 2003; Tsiros and Dimopoulos, 2003), still much remains to be learned about the mechanisms of Hg (II) partitioning between solid and liquid phases. The processes occurring in the soil are particularly complex and poorly understood (Scholtz et al., 2003). A further understanding of the mechanisms of Hg (II) adsorption and desorption reactions is urgently needed to improve model predictions.

Due to the high susceptibility of Hg2+ to form complexes, only a small fraction of this ion occurs in soil solution. A major fraction of Hg2+ is either bound to soil minerals or adsorbed onto solid inorganic and organic surfaces (Steinnes, 1995). Many environmental factors can interfere with the Hg adsorption–desorption process. Factors identified as important for the fate and behavior of Hg (II) from soils include: Hg speciation, soil pH, chloride ions, organic matter content, form and content of soil colloids, competitive inorganic ions, and so on (Schlüter, 2000; Grigal, 2002). A large number of reports have shown that pH is the single most important parameter affecting the adsorption of Hg (II) onto goethite. It has been demonstrated that the pH50% (defined as the pH at which 50% of added metal is adsorbed) strongly correlates with the first hydrolysis constant of the metal (Tiller et al., 1984). It is a well known fact that Hg2+ is strongly bound to the organic matter present in soil (Yin et al., 1997; Grigal, 2003), indicating that under these circumstances the metal has low mobility, especially under low pH conditions (Yin et al., 1996). In addition, the sorption of Hg2+ onto particle surfaces can be significantly affected by the presence of complexing ligands. Inorganic ligands include chloride, phosphate, sulfate, and sulfide, and organic ligands such as low molecular weight organic acids (LMWOA, amino acid, citric acid etc.), fulvic, and humic acids are also common. Most ligands can lower or enhance the adsorption of metal cations due to several possible processes including: (i) formation of stable nonadsorbing metal–ligand aqueous complexes; (ii) formation of metal–ligand ternary surface complexes, which can lead to surface precipitation at high metal and ligand concentrations; (iii) competitive ligand sorption to particle surfaces, effectively blocking reactive adsorbing sites on the surface; and (iv) reduction of the positive charge on clay mineral surfaces (assuming that ligands are anions and pH is below the pHpzc of the clay minerals), thereby lowering the electrostatic repulsion of cations by the surfaces (Kim et al., 2004).

Understanding surface sequestering processes in soils should allow us to better evaluate the bioavailability and, hence, potential toxicity of trace metals to organisms, including human beings (Appel and Ma, 2002). However, minimal information is available on the characteristics of Hg adsorption–desorption in paddy soils, which are widespread in China. The overall objectives of this research were to investigate the adsorption–desorption of Hg in two paddy soils and to understand the mechanisms involved in the surface reactions.


   Materials and Methods
TOP
 NOTES
 ABSTRACT
 INTRODUCTION
 Materials and Methods
Results
 Discussion
 Conclusions
 REFERENCES
 
Soil Samples
Two paddy soils used in this study were collected at 0- to 20-cm depth in Zhejiang Province, southeastern China: a silty loam soil (SLS soil) from Huajiachi campus, Zhejiang University, Hangzhou, and a yellowish red soil (YRS soil) from Deqing County, Zhejiang province. Both belong to Gleyi-Stagnic Anthrosols in FAO/UNESCO nomenclature. After collection, composite samples of the soils were air-dried, ground, and passed through a 2-mm sieve before use. Some basic physicochemical properties of the soils are listed in Table 1 . Soil pH was measured using a pH meter (Thermo Orion 250, Orion Research, Inc., Boston, MA, USA) at a soil/solution ratio of 1:2 in either deionized water or 1 mol L–1 KCl. Soil organic carbon was determined by the modified Tinsley method (Tinsley, 1950). Particle size distribution was measured by the hydrometer method (Liu et al., 1996). Cation exchange capacity (CEC) was determined using 1 mol L–1 NH4OAC following the procedure described by Bao (Bao, 1999). Exchangeable Hg was extracted by 0.1 mol L–1 HCl at a soil/solution ratio of 1:5 (Sparks, 1996), and total Hg in the soil sample was determined by the aqua regia digestion method (Sparks, 1996). The concentrations of Hg in the extracts or digesters were measured using an atomic fluorescence spectrometer (AFS-230E, Beijing Haiguang Instrument Co., Beijing, China). For the metals present in high concentrations the supernatant solution was diluted with deionized water and the concentrations were obtained directly from appropriate calibration curves prepared with the components of the extraction solution diluted by the same factor. The USEPA Method 1631 (USEPA, 1996) was followed for sample handling, preparation (including preservation and storage), and analysis. The water and aqueous extract samples were acidified (pH < 2) using concentrated HCl before analysis. All the samples were stored in fluoropolymer bottle with secured cap in a refrigerator (<4°C) to prevent oxidation. The instrument (AFS) could achieve a detection limit of 0.4 ng/L for Hg in water samples, and a recovery of 75 to 125% for matrix-spiked samples.


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  		Table 1. Basic properties of the tested soil.{dagger}

 
Adsorption of Hg2+ Ion
Portions of 2.0 g air-dried soil were placed into 100-mL polypropylene centrifuge tubes, and 50 mL of 0.01 mol L–1 KCl (pH 5.4) solution containing 0.00, 0.12, 0.24, 0.40, 0.60, 0.80, 2.00, 8.00, 20.00, and 40.00 mg Hg L–1 (as HgCl2) were added to each tube. The suspensions were shaken at 200 rpm for 2.0 h at 25°C and then equilibrated in a dark incubator for an additional 22 h, a time previously found to be sufficient for equilibration. No pH control was imposed. At the end of the designated time, the suspensions were centrifuged at 2000 x g relative centrifugal force for 10 min and filtered. Ten milliliters of the filtrate were transferred into a 10-mL polypropylene centrifuge tube for measuring Hg2+ concentration using the AFS. Total amounts of adsorbed Hg2+ were calculated by the difference between the total applied Hg2+ and the solution Hg2+ in the equilibrium solution. The remaining solution was used for measuring pH.

Desorption of Adsorbed Hg2+ Ion
The tube with the soil residue separated from the supernatant solution by centrifugation was weighed to measure the residual Hg2+ in the solution. Fifty milliliters of 0.01 mol L–1 KCl (pH 5.4) were added to each tube containing the Hg-enriched soil residue. The suspensions were shaken at 200 rpm for 2 h at 25°C and equilibrated for an additional 22 h. The equilibrated suspensions were then centrifuged at 2000 x g relative centrifugal force for 10 min and filtered. Ten milliliters of the filtrate were transferred into a 10-mL polypropylene centrifuge tube for measuring Hg2+ concentration. The remaining solution was used for measuring pH. To estimate the affinity of Hg2+ in soils, the desorption process was repeated five times (D1 to D5). The nonextractable fraction of the adsorbed Hg2+ was obtained by the difference between the total adsorbed Hg2+ and the total recovered Hg2+ by five successive extractions with the KCl solution (pH 5.4) (Yu et al., 2002). All glassware and plastic-ware used in this study were previously soaked in 14% HNO3 (v/v) and rinsed with deionized water. All reagents used were of analytical grade or better.

Statistical Analysis
All data were processed by Microsoft Excel, and the regression was conducted using the programs of Statistical Package SPSS 10.0 (ZheJiang University) (SPSS Inc., 1999).


   Results
TOP
 NOTES
 ABSTRACT
 INTRODUCTION
 Materials and Methods
 Results
Discussion
 Conclusions
 REFERENCES
 
Adsorption Isotherms of Hg2+ Ion
Adsorption isotherms of Hg2+ were constructed to compare metal adsorption capacity between the two paddy soils. The adsorption characteristics of Hg2+ in the two paddy soils were similar. Mercury adsorption was greater in the YRS than the SLS soil. Mercury applied at 0 to 10 mg kg–1 was mostly adsorbed in both soils. Adsorption of Hg2+ increased steeply with Hg2+ concentration in the equilibrium solution at low concentrations (10–200 mg kg–1) for SLS soils and (10–500 mg kg–1) YRS soil, and the increase was diminished at the equilibrium Hg2+ concentrations >200 mg kg–1 for SLS soil and 500 mg kg–1 for YRS soil (Fig. 1 ). The YRS soil adsorbed more Hg2+ than the SLS soil at the same Hg2+ equilibrium concentrations. At the highest level of added Hg2+ (1000 mg kg–1), the YRS soil adsorbed 37.4% of the applied Hg2+, as compared with 27.4% for the SLS soil (Table 2 ), probably due to its higher organic matter and clay content, and larger external surface area (Table 1), thus adsorbing more Hg2+ than the SLS soil.


Figure 1
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  		Fig. 1. Isotherms of Hg2+ adsorption in the paddy soils. Data are means of three replications. Error bar indicates standard error.

 

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  		Table 2. The adsorption and desorption of Hg2+ in two paddy soils.

 
Mercury adsorption in both soils was well described by the Freundlich (logX = logKF +1/nlogC) and the simple Langmuir (C/X = C/Xm+1/(XmK)) adsorption models. The correlation coefficients (r2) were 0.987, P < 0.01 for SLS soil and 0.979, P < 0.01 for YRS soils with the Freundlich model, but were 0.999, P < 0.01 for SLS soil and 0.999, P < 0.01 for YRS soil with the Langmuir model. The Langmuir model gave a better fit with the adsorption data than the Freundlich model. The result was not consistent with the conclusion of Carey et al. (1998) that the Freundlich model better describes the adsorption behavior of Hg2+ removal than the Langmuir model. The monolayer maximum adsorption (Xm) from the Langmuir equation is frequently used for comparing potential adsorption capacity of different soils and soil components. The Xm value was 111 mg kg–1 for the SLS soil and 213 mg kg–1 for the YRS soil. The physical meaning of K from the Langmuir equation is not well defined. However, it is usually considered to relate the binding energy of Hg2+ adsorption. The greater the K value is, the more tightly the adsorbed Hg2+ is bonded. The SLS soil, though with a smaller adsorption capacity, had a greater K value than the YRS soil (Table 3 ). The product of Xm and K (MBC = XmxK) from the Langmuir equation reflects the maximum buffer capacity of the soil for Hg2+. The value of MBC was 25.6 for the SLS and 25.5 for YRS soil, suggesting that the two soils had a similar buffering capacity for Hg2+. A similar trend of Hg (II) adsorption in soil was reported by Cruz-Guzmán et al. (2003) and Gupta et al. (2004).


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  		Table 3. Isothermal characteristics of Hg2+ adsorption in the tested soils.

 
Effect of Hg2+ Adsorption–Desorption on Soil pH
The pH of equilibrium solution decreased with Hg2+ adsorption for both soils (Table 4 ). The pH of YRS soil decreased more than the SLS soil at the same Hg2+ loading. For the same soil, equilibrium solution pH decreased with increasing Hg2+ adsorption. The maximum pH drop was up to 0.75 units for the SLS soil and 0.91 units for the YRS soil. Obviously, H+ and/or Al3+ were released during Hg2+ adsorption. These results suggest that severe Hg2+ contamination potentially causes soil acidification, and the more Hg2+ is loaded, the more acute the acidification may be. The decrease in soil pH was quadratically correlated with the amounts of Hg2+ adsorbed for both the SLS and YRS soil (r2 = 0.977, 0.986, P < 0.01 for SLS soil and YRS soil, respectively).


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  		Table 4. Equilibrium solution pH in relation to Hg2+ adsorption–desorption.{dagger}

 
After five successive desorptions, equilibrium solution pH increased, probably because of H+ retention during Hg2+ desorption. Equilibrium solution pH generally increased with desorption and was lower when Hg2+ concentration in the equilibrium solution increased (Table 4). However, the equilibrium solution pH was lower than the original at high initial Hg2+ concentrations, possibly because of some residual Hg2+ that was not desorbable. A greater decrease in equilibrium solution pH was observed in the YRS soil than the SLS soil at the same Hg2+ loading, because of more Hg2+ being adsorbed in the YRS soil than the SLS (Table 2).

The relationship between pH decrease ({Delta}pH) and the amounts of Hg2+ adsorbed fitted well a multinomial regression equation (y = ax2+bx+c, where y is {Delta}pH, x is the amount of Hg2+ adsorbed, and a, b, and c are constants), with correlation coefficients R2 = 0.98 and 0.94, P < 0.01, respectively, for the SLS and YRS soils (Fig. 2 ). At higher levels of adsorbed Hg2+, the pH decrease became less in the two soils, probably because of approaching maximum adsorption and limited availability of H+ (Table 4). The pH–Hg2+ adsorption relationship from this study is consistent with previous findings by Yang et al. (2004) and He et al. (2005).


Figure 2
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  		Fig. 2. The pH changes in relation to Hg2+ adsorption in the two paddy soils. Error bar indicates standard error.

 
Distribution Coefficients of Hg2+
The distribution coefficient is defined as the ratio of adsorbed Hg2+ to dissolved Hg2+, i.e., the ratio of Hg2+ in solid to liquid phase. This parameter reflects the affinity of Hg2+ to soil surface (Wang, 1999). In this study, Hg2+ was almost completely adsorbed at low initial concentrations (<15 mg kg–1) and Hg2+ concentration in the equilibrium solution was below detection limits (≤0.005 ng mL–1). As a result, the Kd values were not obtainable at these low Hg2+ loadings. The Kd values were high at low Hg2+ loadings (<15 mg kg–1), and decreased rapidly with increasing external Hg2+ loadings from 15 to 200 mg kg–1 for both soils, but the decrease became slow at higher Hg2+ loadings (>200 mg kg–1). This may be attributable to the high affinity of Hg2+ to some highly selective sites at low concentrations and low affinity for those less selective sites at high Hg2+ concentrations (Basta and Tabatabai, 1992a). The YRS soil had a much higher Kd value than the SLS soil (Fig. 3 ). The difference may be attributed to the higher organic matter and clay contents of the YRS soil, as compared with the SLS soil (Table 1).


Figure 3
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  		Fig. 3. Effect of initial Hg2+ concentration on Hg2+ distribution coefficient (Kd) in the two paddy soils. Data are means of three replications.

 
The reaction of heavy metal adsorption on soils can be universally described as:

Formula 1[1]
The linear form of Eq. [1] is expressed as:

Formula 2[2]
where Kd is the distribution coefficient; K' = Kxax[S(OH)n] is a constant; and n is the average number of H+ released by adsorbing one Hg2+.

There was a linear relationship between the Kd value and pH of adsorption equilibrium solution for both soils (r2 = 0.978 and 0.986, P < 0.01 for the SLS and YRS soil, respectively) (Fig. 4 ). For the paddy soils, H+ is released mainly through specific adsorption of Hg2+, with a small contribution from nonspecific adsorption process. Therefore, the adsorption of Hg2+ on the paddy soil surface involves mainly specific adsorption, with a small portion of nonspecific adsorption. In this study, a greater n value was observed with the YRS soil (2.84) than the SLS soil (2.80). This agreed with the previous finding that the adsorption equilibrium solution pH of the YRS soil was much lower than that of the SLS soil and it decreased more rapidly with the amount of Hg2+ adsorbed.


Figure 4
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  		Fig. 4. The relationship between Kd and pH of adsorption equilibrium. Data are means of three replications.

 
Desorption of Adsorbed Hg2+
Desorption of the adsorbed Hg2+ was very small in the 0.01 mol L–1 KCl (Table 2). No desorption occurred at the adsorbed Hg2+ up to 15 mg kg–1. Desorption increased with increasing Hg2+ adsorption saturation for both soils, though with a lower rate. The YRS soil desorbed less Hg2+ than the SLS soil at the same Hg2+ concentrations (Figs. 5 and 6 ). After five successive desorptions, the maximum amounts of Hg2+ desorbed accounted for only 24.4% of the adsorbed Hg2+ for the SLS soil and 14.4% for the YRS soil (Table 2). The YRS soil that had a greater adsorption capacity desorbed less Hg2+ than the SLS soil at the same amount of adsorbed Hg2+.


Figure 5
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  		Fig. 5. Fractions of five successive extractions and residual Hg2+ in the total adsorbed Hg2+ in the SLS soil. The terms D1 to D5 represent the fraction of desorbed Hg2+ from each of the five successive extractions and the residual Hg2+ represents the fraction of adsorbed Hg2+ that was not recovered by the five successive extraction with 0.01 mol L–1 KCl (pH 5.4). Data are means of three replications.

 

Figure 6
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  		Fig. 6. Fractions of five successive extractions and residual Hg2+ in the total adsorbed Hg2+ in the YRS soil. The terms D1 to D5 represent the fraction of desorbed Hg2+ from each of the five successive extractions and the residual Hg2+ represents the fraction of adsorbed Hg2+ that was not recovered by the five successive extraction with 0.01 mol L–1 KCl (pH 5.4). Data are means of three replications.

 
The proportion of the adsorbed Hg2+ that was not desorbed by the KCl decreased with the increasing concentration of adsorbed Hg2+ (Fig. 7 ). The portion of Hg2+ that was adsorbed but not desorbed by the five successive desorptions was likely related to high binding energy, and may not be available to plants. There was a quadratic relationship between the amount of Hg2+ desorbed and the amount of adsorbed for the two soils (RSLS2 = 0.989, P < 0.01; RYRS2 = 0.885, P < 0.05). Both the SLS and YRS soils can retain a large amount of applied Hg2+. This portion of residual Hg2+ after five successive desorptions measures the potential fixation of Hg2+ by the soils.


Figure 7
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  		Fig. 7. The relationship between Hg2+ adsorption and desorption. Data are means of three replications.

 
The data presented in Table 2 indicate that not all adsorbed Hg2+ can be readily desorbed (Schultz et al., 1987). Similar results have been found using batch techniques (Yin et al., 1996). These results suggest that the adsorption and desorption processes are not reversible. In other words, there is hysteresis between adsorption and desorption of Hg2+ in soil. Yin et al. (1997) investigated Hg (II) desorption from four different types of soils in a stirred flow reaction chamber and also found that adsorption–desorption hysteresis was evident.

The hysteresis of Hg(II) adsorption/desorption is related to the rate and mechanisms of Hg(II) retention to and release from soil. Some researchers suggested that hysteresis resulted from the binding of metals to different sites. Amacher et al. (1990) proposed a multi-reaction model for describing the reactions of several metals with soil. They assumed that adsorption/desorption processes involved three concurrent reactions with three different kinds of sites on soil, i.e., a rapid and reversible reaction, a slow and reversible reaction, and an irreversible reaction.

The principal causes of the observed irreversibility may include the stability of surface Hg2+ complexes and the mechanisms through which the adsorption occurs. The strong binding of Hg(II) to high affinity sites on soil organic matter such as S-containing groups (Reimers and Krenkel, 1974), and diffusion through intra-particle micropores of soil organic matter seems to be responsible for the extent of hysteresis. The latter process has often been found to be the rate-limiting step (Yin et al., 1997). Therefore, a higher organic matter content in soil often results in slow reaction rates and a higher proportion of residual Hg (II) after desorption (Bouchard et al., 1988).


   Discussion
TOP
 NOTES
 ABSTRACT
 INTRODUCTION
 Materials and Methods
 Results
 Discussion
Conclusions
 REFERENCES
 
The different characteristics of Hg2+ adsorption–desorption between the two paddy soils may be attributed to the different physicochemical properties of the soils. There were higher organic C and clay contents in the YRS soil, as compared with the SLS soil (Table 1). Organic matter is the main organic component in soil and humus is a sustainable source of chelating agents. Generally, natural organic matter (NOM), such as humic and fulvic acids, contain a large number of functional moieties such as carboxylic, phenolic, and alcoholic groups that interact with clay mineral surface groups as well as ions in solution. Studies showed that humic acid (HA) can form stable complexes with different heavy metal ions, including Hg (II) (Mierle and Ingram, 1991). Miretzky et al. (2005) investigated the sorption of Hg (II) onto four different types of Amazon soils from the A-horizon by means of column experiments under saturation conditions and controlled metal loads and suggested that higher organic matter content in the soil resulted in higher Hg (II) adsorption. Similar results, in which the soil adsorption capacity of Hg (II) correlates positively with the organic carbon content, were reported by Yin et al. (1996), Grigal (2003), and Schuster (1991). Clay minerals and metal oxides, as main active soil colloidal constituents, are important sorbents of heavy metals, owing primarily to their cation-exchange capacity (CEC) and their ability to form inner-sphere complexes through surface reactive groups such as OH (Ponizovskii and Mironenko, 2001). Mutual interactions between clay minerals, Fe oxides, and organic matter can greatly alter the sorptive properties of these soil constituents for heavy metals because such interactions usually involve cation exchange sites and carboxyl and OH surface groups, that is, potential sorption sites for heavy metals. Traditionally, the importance of the different soil components in contaminant sorption is evaluated by determining the sorption behavior of selected soil components or by investigating changes in sorption after removing selected soil constituents (Li et al., 2001).

Several studies on metal adsorption by soils have demonstrated close relationships between metal adsorption and soil pH or CEC (Basta and Tabatabai, 1993). Soil pH may affect metal hydrolysis, ion-pair formation, organic matter solubility, as well as surface charge of iron and aluminum oxides, organic matter, and clay edges (Sauve et al., 1998). Raising soil pH increases cationic heavy metal retention to soil surfaces via physical sorption, inner-sphere surface complexation, and/or precipitation and multinuclear type reactions (McBride, 1994). Several investigators have especially documented the influence of pH on the adsorption of Hg (II) by a variety of mineral surfaces. MacNaughton and James (1974) studied the adsorption of Hg (II) by SiO2 and observed a sharp increase in Hg (II) adsorption between pH 2 and 3, with an adsorption maximum at a pH of 4. Lockwood and Chen (1974) studied the adsorption of Hg (II) by Fe(OH)3 and also observed a sharp increase in retention in the pH 3 to 4 range. They concluded that Hg(OH)20 was the major Hg (II) species adsorbed. Numerous other investigators have observed the abrupt increase in Hg (II) retention between pH 2 and 4, with adsorption maxima between pH 4 and 5, and a shift in adsorption maxima to higher pH values when Cl–1 is present, irrespective of adsorbent type (Yin et al., 1996). These studies have also found that Hg (II) adsorption is strongly correlated to the formation of the hydrolysis products HgOH+ and Hg(OH)20, leading to the conclusion that these aqueous species are preferentially adsorbed by mineral surface. Some studies have shown that CEC is the predominant property affecting adsorption of Pb (Soldatini et al., 1976), Cd (Hassett, 1974), and Zn (Shuman, 1975). Previous studies also showed that Zn and Pb adsorption by soil is closely related to CEC and pH (Zimdahl and Skogerboe, 1977). Different mechanisms might be involved in Hg2+ adsorption/desorption at different levels of Hg2+ loading and between the two soils. The YRS soil may have more adsorption sites and form much stronger surface complexes with Hg2+ than the SLS soil. Perhaps Hg2+ adsorption sites are different between the two soils. Mercury adsorbed by the SLS soil may be mainly affected through cation exchange and diffusion into "dead-end" or micropores created by crystal defects, but the formation of inner-sphere complexes may be the predominant mechanism of Hg2+ adsorption in the YRS soil and be less readily desorbed. Additionally, iron and aluminum oxides may contribute to the high affinity of Hg2+ by the YRS soil due to its larger amount of free Fe oxides than SLS soil.

Several principally different processes influence the mobility of Hg2+ ions in the aquatic and the geological environment. These processes range from surface reactions, i.e., sorption and desorption, surface complexation with inorganic and organic ions such as chloride and fulvic acid, reduction and oxidation, and exchange at the soil/air interface (Xu and Allard, 1991; Gunneriusson and Sjöberg, 1993). Like other trace metals, the fate and behavior of Hg(II) in the environment is largely controlled by the degree and mechanism of Hg (II) adsorption and desorption reactions with various adsorbents. Adsorption processes are of great importance in the regulation of Hg. Many batch experiments for the sorption of Hg(II) have been reported, and different models have been successively used to describe these experiments (Stumm, 1992; Toulhoat et al., 1996). One mechanism of the formation of such organomineral complexes is adsorption of the positively charged metal cations of organic complexes to the negatively charged surface of clay minerals, which leads to immobilization of the metal. It has been recently reported that humic acids are increasingly sorbed to mineral surfaces such as iron oxides with decreasing pH (Avena et al., 1999).

There are still different opinions regarding Hg2+ adsorption models such as hydrolyzed/unhydrolyzed adsorption and monodentate/bidentate reactions. Carey et al. (1998) developed a mass transfer model involving surface equilibrium to simulate Hg(II) removal in a fixed bed. They concluded that the Freundlich model provides a slightly better fit than the Langmuir model during the simulations. Meserole et al. (1999) presented a theoretical model that combines the adsorption characteristics measured in the laboratory with mass transfer considerations to predict Hg(II) removal. In their study, the Freundlich and the Langmuir models were used to describe the surface equilibrium. However, our research suggested that the Langmuir model gave a better fit with the adsorption data than the Freundlich model. Many questions remain to be answered about Hg2+ adsorption on paddy soils. Further studies, especially molecular studies, are needed to identify the mechanism of Hg2+ adsorption/desorption in paddy soils.

A number of factors might contribute to the decreased adsorption equilibrium pH and the difference between the two soils. First of all, more H+ is exchanged by the increased Hg2+ concentrations. Second, as Basta and Tabatabai (Basta and Tabatabai, 1992b) inferred in their article, at low heavy metal loadings, heavy metals might replace adsorbed Ca2+ and Mg2+ because they have less affinity to soil constituents than Al3+. However, at higher heavy metal loadings, exchange reactions between heavy metals and Al3+ might happen, followed by hydrolysis of Al3+ that decreases solution pH, especially in soils that have significant amounts of exchangeable acidity and exchangeable Al3+. Third, the YRS soil had lower pH. As a result, the YRS soil had more hydroxylated surface and subsequently released more H+ at the same amount of Hg2+ adsorbed than the SLS soil. Even in contaminated soils, most of the Hg2+ present as insoluble forms precipitated or bound to the soil surfaces. In our study, Hg2+ was almost completely adsorbed in both soils at 0 to 15 mg kg–1, indicating that the paddy soil had high fixation ability for Hg2+, and the bioavailability of Hg2+ should be low in the paddy soils. However, recent study indicates that although the heavy metal Hg2+ was tightly adsorbed in the soil with very limited mobility, its activation may be caused by increased Hg2+ inputs (Kalbitz and Wennich, 1998).


   Conclusions
TOP
 NOTES
 ABSTRACT
 INTRODUCTION
 Materials and Methods
 Results
 Discussion
 Conclusions
REFERENCES
 
The adsorption and desorption behavior of Hg2+ was well described by the Langmuir and the Freundlich models with correlation coefficients r2 > 0.979. Behavior of Hg2+ adsorption–desorption was related to soil properties. The YRS soil that contained greater amounts of extractable Fe oxides, organic C, and clay, and with a lower pH, had a greater adsorption capacity and affinity for Hg2+ than the SLS soil. The adsorption of Hg2+ caused soil acidification. Equilibrium solution pH decreased with increasing Hg2+ adsorption. The pH decrease was greater in the YRS soil than in the SLS soil. Desorption of the adsorbed Hg2+ was very small in the 0.01 mol L–1 KCl. No desorption occurred at the adsorbed Hg2+ up to 15 mg kg–1. After five successive desorptions, the maximum amounts of Hg2+ desorbed accounted for only 24.4% of the adsorbed Hg2+ for the SLS soil and 14.4% for the YRS soil. These results indicate that adsorption–desorption of Hg2+ in the paddy soils is not reversible, as a major proportion of the adsorbed Hg2+ is not readily desorbed in soil solution. However, the tightly-bonded Hg2+ can be activated and transferred into the environment (atmosphere, food chain, and surface water bodies) under favorable conditions. Understanding adsorption and desorption behavior and transfer pathways of Hg2+ in soils is a key to providing a basis for developing effective measures to prevent soil Hg from entering the food chain and dispersing into the environment.


   ACKNOWLEDGMENTS
 
This study was, in part, supported by a grant from the Natural Science Foundation of China (# 20577044) and a grant from the Science and Technology Ministry of China (# 2002CB410800), and by a Program Funding for Changjiang Scholars and Innovative Research Team of Higher Education of China (#IRT0536).


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 Results
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