Published online 20 November 2006
Published in Vadose Zone J 5:1216-1221 (2006)
DOI: 10.2136/vzj2006.0037
© 2006 Soil Science Society of America
677 S. Segoe Rd., Madison, WI 53711 USA
ORIGINAL RESEARCH
Modeling Dynamic Sorption of Cadmium in Cropland Soils
Weiping Chen*,
Andrew C. Chang,
Laosheng Wu and
Albert L. Page
Department of Environmental Sciences, Univ. of California, Riverside, CA 92521
* Corresponding author (chenweip{at}yahoo.com.cn)
Received 6 March 2006.
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ABSTRACT
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In sorption experiments, Cd in the solution phase may be surface adsorbed or immobilized and precipitated into mineral phases. The reaction kinetics can be described by a two-site model combining a linear instantaneous model for the surface adsorption and a first-order reaction kinetic model with forward and backward reaction constants for the immobilization and precipitation of the mineral phase. A simplified sequential procedure was developed to study the contribution of these two processes. Results of batch Cd adsorption experiments with two California soils were used to illustrate the model validity. When the Cd sorption was obtained by varying the initial Cd solution concentration and maintaining constant equilibration time, the amounts of Cd in the solid phase as well as the adsorbed and mineral phases increase linearly with the initial Cd solution concentrations. As much as 90% of the sorbed Cd may be in the mineral phase. When the Cd sorption was examined at varying equilibration time and a constant Cd initial solution concentration, the sorbed Cd in the mineral phase increased exponentially to approach a maximum with time. The forward and backward reaction constants were obtained by fitting the sorption data to the model equations. It showed that the forward reaction was five times faster than the backward reaction for both soils and the reaction rates were two times faster in the heavier textured Holtville clay loam soil than the Arlington sandy loam soil.
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INTRODUCTION
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CADMIUM MAY BE INTRODUCED into cropland soils through a variety of agricultural production related activities such as fertilizer applications, irrigation, pesticide sprays, and waste disposal (Chang and Page, 2000). Krage (2002) showed that the As, Cd, and Pb contents of some irrigated cropland soils in California had become elevated in comparison with the baseline levels. This enrichment in the soil could lead to inadvertent transfer of trace elements through the food chain to unsuspecting consumers. The possible risks from the increasing amount of Cd in cropland soils have resulted in growing public health concerns of consumers (Krishnamurti et al., 1999; de Meeus et al., 2002). It is imperative to understand the behavior of Cd in soils.
In soil, Cd is adsorbed onto the surface of hydrous oxides of Fe and Mn, organic matter, and clays, forms inorganic precipitates, and may be present as primary and secondary minerals (Gray et al., 2000). The behavior of Cd in the soil–water–plant system is dependent on its distribution among these forms and the chemical reactions between the forms. The Cd present in soil solution, the soluble form, is therefore of special interest as it is readily absorbed by plants and is mobile in the vadose zone. The potential for soil-borne Cd to cause environmental and health problems is closely associated with its mobility and bioavailability, which in turn are controlled by the chemical interactions taking place on the solution–solid interface. Knowledge of the sorption of Cd by soil is essential for assessing environmental and health risks associated with soil Cd of anthropogenic origins.
The sorption processes of Cd in soils have been extensively studied (Sanchez-Martin and Sanchez-Camazano, 1993; Salim et al., 1996; Gao et al., 1997; Naidu et al., 1997; Filius et al., 1998; Pardo, 2000; Li et al., 2001). Frequently, the linear or Freundlich isotherm model is used to depict the equilibrium between Cd in the solution and solid phases and in turn predict the retention of Cd by soils and its transport through soil columns (Boekhold and van der Zee, 1991; Palm, 1994; Tiktak et al., 1998; Keller et al., 2001; Senutjens et al., 2002); however, the models based on solution–solid phase equilibrium are not always adequate to describe outcomes that are the result of several reactions occurring in soil at the same time (Barrow, 1989). Kinetic reaction models may describe the time-dependent solute sorption and desorption (Ma and Selim, 1994; Selim et al., 1992b, 1999). Most commonly, the model is based on first-order reversible reaction kinetics. Amacher et al. (1986) noted that first-order reaction kinetics were not fully capable of describing the retention of Cd, Cr, and Hg in soil over time.
The chemical reactions involving Cd in the solution–solid phase are complex, and different reaction mechanisms are often involved such as ion exchange, coprecipitation with soil minerals, and the formation of complexes with soil organic matter. There are different ways of looking at these processes individually. For modeling purposes, it is necessary to simplify and look these processes as a whole. Based on the affinity of the sorption sites, some multireaction models have been introduced (Brusseau et al., 1989, 1992). The multireaction model was successful in describing the retention behavior of soil Cd, Cr(VI), and Hg (Amacher et al., 1988, 1990) over time; however, those approaches tend to overparameterize, making application to field situations difficult (Selim et al., 1992b).
In this study, batch adsorption experiments of two California cropland soils at fixed equilibration time and at fixed initial solution concentration were conducted. The dynamic sorption was characterized in a two-site model by combining a linear equilibrium reaction model and a first-order reaction kinetics model. The model parameters were obtained by fitting the experimental data to the model.
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MATERIALS AND METHODS
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Conceptual Considerations
The sorption of Cd from solution by soil components usually consists of two steps: an initial rapid adsorption followed by a slow removal process. The initial rapid adsorption is generally ascribed to reversible adsorption at the interface. The characteristic time scale of trace element adsorption on the soil surface is typically on the order of seconds to minutes, and therefore is commonly quantified by thermodynamic equilibrium distribution models. The slow removal process may be due to surface precipitation, coprecipitation, or further diffusion into inner sites of soil particles, which is on a time scale of hours to days and may be described by a first-order kinetic reaction. Upon entering soil, the Cd undergoes these two types of reactions simultaneously. The distributions of Cd between the solution, surface adsorbed, and immobilized mineral phases are dependent on the Cd concentration of the solution phase at the start, C0, and the time to equilibrium, t. In the two-site model, the different sorption reactions are ascribed as either surface adsorbed, which is instantaneously in equilibrium with the Cd in the solution phase, or as occluded or coprecipitated with soil minerals, which follows a first-order kinetic equilibrium with the solution phase or adsorbed phase. Consequentially, the Cd held on those solid soil components are defined as Cd in the adsorbed phase and mineral phase, respectively.
The mass balance in a Cd sorption experiment when these two processes occur simultaneously can be described as follows:
 | [1] |
where R is the water to soil ratio (L kg–1), ka is the linear adsorption coefficient for Cd distribution between the solution and adsorbed phase (L kg–1); C0 is the initial Cd concentration of the solution phase (µg L–1); and Ct (µg L–1) and MPt (µg kg–1) are the Cd concentrations in the solution phase and immobilized mineral phase, respectively, at a specific equilibration time (t). The slower reaction kinetics of the immobilized mineral phase can be described by the following first-order reaction equation:
 | [2] |
where kf and kb are the forward and backward reaction rate constants (h–1) corresponding to the immobilization and dissolution processes, respectively, and Kd is the partitioning coefficient. Combining Eq. [1] and [2], the time-dependent change of the Cd in the solution phase and mineral phase is given by
 | [3] |
 | [4] |
If t is fixed, the right-hand sides of Eq. [3] and [4] may be reduced to linear forms in terms of C0. The Cd concentrations in the solution phase (Ct) and mineral phase (MPt) at a specific t may be obtained for various C0, and Ct vs. t would be a straight line. When C0 is fixed, the right-hand sides of Eq. [3] and [4] may be reduced to an exponential form. The value of Ct will decrease exponentially with respect to equilibration time to a minimum, and MPt will increase exponentially with respect to time to a maximum.
Batch Cd adsorption experiments may be conducted and the data may be fitted into Eq. [3] and [4] to evaluate the sorption kinetics, namely ka, kf, and kb.
Soil Materials
Two soils, an Arlington sandy loam soil (coarse-loamy, mixed, active, thermic Haplic Durixeralf) obtained at the University of California Riverside Agricultural Experiment Station and a Holtville clay loam soil (clayey over loamy, smectitic over mixed, superactive, calcareous, hyperthermic Typic Torrifluvent) obtained at the University of California Meloland Field Station in Imperial Valley, were used in the batch Cd adsorption experiments. These two soils were selected to illustrate the applicability of our approach to soils of divergent characteristics. Table 1 presents the general properties of these two soils. Except for the total Cd content, the remainder of the parameters were based on data as reported by the National Cooperative Soil Survey (2006). The cation exchange capacity (CEC) was determined by NH4OAc extraction at pH 7. This method inherently would underestimate the CEC of calcareous soils such as the Holtville loam. The total Cd contents were analyzed by graphite furnace atomic absorption spectroscopy (AAS-GF) after dissolution of soils based on the standard USEPA 3052 microdigest method (USEPA, 1996).
Determination of Cadmium in Solid Phases
Equations [3] and [4] require estimates of Cd in adsorbed (kaCt) and mineral (MPt) phases. The distribution of Cd in the solid phases can be examined with a sequential extraction procedure. Although the sequential extraction may have procedural shortcomings, it is a practical tool to probe the behavior of trace elements such as Cd that are present in soils mostly in the solid phases. Many sequential extraction procedures have been developed to evaluate the chemical forms and distributions of Cd in different background matrices (Soon and Bates, 1982; Kuo et al., 1983; Miller and McFee, 1983; Xian, 1987, 1989; Ramos et al., 1994; Chlopecka et al., 1996; Sanchez et al., 1999; Ahnstrom and Parker, 1999).
Besides for the adsorbed and mineral phases defined above, Cd may form higher stability chelate complexes with soil organic matter that consists of a mixture of plant and animal products in various stages of decomposition. The decomposition of organic matter may release Cd into the solution phase; however, the kinetics is generally on a scale of months to years and the organic matter content only accounts for a small proportion of a soil, thus the contribution from the mineralization of organically complexed Cd may be negligible in a short time frame.
Based on outcomes reported in the literature and the results of preliminary experiments on materials of known composition, a sequential extraction procedure was devised to separate Cd in the soils into the adsorbed and mineral phases. In the sequential extraction procedure, the soil sample was first reacted with 20 mL of 1 M CaCl2 at pH 7.0 in a reciprocating shaker for 2 h at room temperature (Young, 2000). The solution and solids were separated by centrifugation. The solution was saved and the recovered solid was treated twice with 10 mL of deionized water, equilibrated for 30 min by shaking, and separated by centrifugation. The adsorbed-phase Cd was obtained by determining the Cd concentration in the combined solutions. The residual solid was further treated with 20 mL of 1 M HNO3 (Soon and Bates, 1982) and equilibrated by shaking for 6 h at room temperature. After centrifugation, the Cd in the solution was determined as the mineral-phase Cd. The recovery rate based on this procedure may not be 100%, especially at higher loading of Cd, but the fraction that is not dissolved in 1 M HNO3 generally is small and has a very low mobility and thus minor effects on the kinetics study.
Batch Cadmium Adsorption Experiments
The first set of batch adsorption experiments was conducted at initial Cd solution concentrations of 50, 100, 200, 400, 600, 800, 1000, 1500, and 2000 µg L–1 with an equilibration time of 48 h. The different levels of Cd solution were obtained by diluting Puro-Graphic calibration standard for atomic absorption (Buck Scientific, 1002 µg mL–1 Cd in 2% HNO3) with 0.01 M NaNO3. The second set of batch adsorption experiments was conducted at an initial Cd solution concentration of 100 µg L–1 and equilibration times of 2, 4, and 8 h, and 1, 2, 4, 6, 8, 10, and 15 d. For both sets of adsorption experiments, a soil/solution ratio of 1:20 (w/v) was used. The mixtures were allowed to equilibrate on a reciprocating shaker at room temperature. No attempt was made to control the pH during the adsorption experiment. The soils were well buffered and the pH of soil suspensions did not change significantly (<0.3 pH unit). The equilibrium pHs for the Holtville clay loam and Arlington sandy loam soils were 8.0 and 6.5, respectively. Following equilibration, samples were centrifuged and the supernatant was decanted and filtered with 0.45-µm hydrophilic PVDF membrane filters. The sample solution is then determined by AAS-GF. The solid phase was recovered and underwent sequential extraction to recover Cd in the adsorbed and mineral phases.
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RESULTS AND DISCUSSION
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Sorption at Fixed Equilibrium Time
Once introduced into the soil, Cd in the solution phase will approach a dynamic equilibrium with that held on the solid soil component. The adsorption isotherms of Cd for the two California cropland soils, along with the contribution of the surface adsorbed and mineral phases, were plotted in Fig. 1
and 2.
The Cd adsorption isotherms are linear for the initial solution concentration ranging from 50 to 2000 µg Cd L–1. The slopes for the Holtville clay loam soil and the Arlington sandy loam soil, i.e., the partitioning coefficients (Kd), are 7760 and 5554 L kg–1, respectively. The Cd adsorptive ability of the Holtville clay loam is greater than that of the Arlington sandy loam soil, which corresponds with its higher clay content and the total cation exchange capacity (see Table 1), which favor Cd sorption.
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Fig. 1. Adsorption isotherms of Cd for the Holtville clay loam soil, equilibrated at 1:20 (w/v) for 48 h at an initial solution concentration ranging from 50 to 2000 µg Cd L–1.
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Fig. 2. Adsorption isotherms of Cd for the Arlington sandy loam, equilibrated at 1:20 (w/v) for 48 h at an initial solution concentration ranging from 50 to 2000 µg Cd L–1.
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The Kd approach is the common way to describe the distribution of trace elements between the solution and solid phase; however, not all of the Cd in the solid phase is available immediately. As shown in Fig. 1 and 2 for both soils, most of the added Cd was present in the mineral phase, which means the release of this portion of Cd follows a first-order kinetic reaction. The Cd in the adsorbed phase only accounts for a small portion. Based on the slope, the linear distribution coefficient (ka) is 71 and 112 L kg–1 for the Holtville and Arlington soils, respectively, which is much less than the commonly used partitioning coefficient (Kd).
The kinetics parameters, kf and kb, cannot be obtained since the equilibration time is fixed. As shown in Eq. [5] and [6], when t is fixed, Eq. [3] and 4 reduce to simple linear forms with zero intercept, which are given by
 | [5] |
 | [6] |
where a1 and a2 are functions of the water to soil ratio used in the sorption experiment (R) and the kinetics parameters ka, kf, and kb. The equilibrium solution Cd concentration (Ct = 48) and the amount of Cd retained in the mineral phase (MPt = 48) increase linearly in proportion to the initial solution concentration (C0). The experimental results agreed with the theoretical prediction (Fig. 3
and 4)
. The reaction kinetics in the Holtville clay loam were slightly faster than in the Arlington sandy loam, in which the equilibrium concentration is lower and much more Cd was retained in the mineral phase.
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Fig. 3. Cadmium adsorption: Retention in the mineral phase (MP) vs. initial solution concentration (C0), equilibrated at 1:20 (w/v) for 48 h.
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Fig. 4. Cadmium adsorption: Equilibrium concentration vs. initial solution concentration, equilibrated at 1:20 (w/v) for 48 h.
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Sorption at Fixed Initial Solution Concentration
To obtain the kinetic parameters, the batch adsorption experiments were conducted at an initial Cd concentration in solution of 100 µg L–1 and equilibrated for periods from 2 h to 15 d. Under the circumstances, the Cd solution concentration decreased exponentially to approach the equilibrium concentration and the amount of Cd in the mineral phase increased exponentially to approach the maximum with the time of equilibration (Fig. 5
and 6)
. The dynamic sorption can be described by Eq. [3] and [4]. When the initial solution concentration is fixed (C0 = 100 µg L–1), Eq. [3] and [4] reduce to exponential forms, which are given by
 | [7] |
 | [8] |
where a3, a4, and a5 are functions of the water to soil ratio used in the sorption experiment (R) and the kinetics parameters ka, kf, and kb, and b is the sum of kf and kb.
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Fig. 5. Cadmium solution concentration at different equilibration times for the Arlington sandy loam and Holtville clay loam, equilibrated at 1:20 (w/v) at initial concentration of 100 µg L–1.
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Fig. 6. Cadmium in the mineral phase of Arlington sandy loam and Holtville clay loam at different equilibration times, equilibrated 1:20 (w/v) at 100 µg L–1.
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The experimental results agree with the conceptual framework defined by the above equations. When the data is fitted into Eq. [7] or [8], the kinetic parameters of the forward and backward reaction rates may be obtained if the linear Cd distribution constant (ka in Eq. [3]) was defined. Based on the model assumption, the ka values are the slope of the linear regression of the Cd in the adsorbed phase vs. the equilibrium concentration (Fig. 7
). The R2 values for the Holtville clay loam and Arlington sandy loam were 0.9934 and 0.9463, respectively. The ka values were 100 and 70 L kg–1 for the Holtville clay loam and the Arlington sandy loam, respectively. Table 2 summarizes the results for the forward and backward reaction rate constants (kf and kb) based on the solution concentration of Cd and Cd contents in the mineral phase.
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Fig. 7. Cadmium in the adsorbed phase vs. the equilibrium concentration in Arlington sandy loam and Holtville clay loam, equilibrated at 1:20 (w/v) at 100 µg L–1.
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Table 2. Fitting results of kinetic adsorption of Cd on the Holtville clay loam and the Arlington sandy loam soils.
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The kf and kb based on the solution concentration data and on the data of the mineral-phase Cd contents are similar. The Cd adsorption experiments that varied the initial Cd solution concentration (fixed equilibration time) and the experiments that varied the equilibration time (fixed initial Cd solution concentration) produced the same outcomes. The precipitation dissolution kinetics of the Cd mineral phase in the Holtville clay loam is approximately twice as fast as that of the Arlington sandy loam. The ratios of kf to kb for the Cd mineral phases are almost the same. For both, the forward reaction (dissolution) rate constant is about five times larger than the backward (precipitation) rate constant. Both sets of data fitted well with the theoretical equations. The solution concentration data are easier to obtain and may be more reliable than the mineral-phase Cd content data, since the latter may be affected by the extraction process of the adsorbed phase; however, the fitted kf and kb are dependent on the value of ka, which in turn is dependent on the extraction data of the adsorbed phase. For example, if ka is 120 L kg–1 instead of 100 L kg–1, the fitted kf and kb will be 0.06 and 0.00024 h–1, respectively. The mineral-phase Cd content data is independent of ka and the values of kf and kb can be obtained directly from the regression model.
The experiments were conducted at a 1:20 soil/water ratio and the mixtures were equilibrated with a reciprocating shaker. Unlike field conditions, the contents were well mixed and the reactions would be more complete. The reaction rate constants, kf and kb, are expected to be less under field conditions.
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CONCLUSIONS
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The developed two-site model, by combining a linear equilibrium reaction model and a first-order reaction kinetics model, can describe the dynamic sorption of Cd in two California soils very well. The model can be further used to predict the fate and transport of trace elements in soils and assess the hazard associated with the introduction of compounds containing trace elements to the soil system (as used in the generalized trace element mass balance model developed by Chen [2005] at the University of California-Riverside). The characterized kinetic parameters can be obtained by conducting the batch adsorption experiments at fixed initial Cd solution concentration and fitting the results to the theoretical equations, based on either the solution concentration data or on the data of the Cd retained in mineral phase. Both approaches fitted well with the theoretical equations and give similar results. The solution concentration data are easier to obtain but the fitted kf and kb are dependent on the fitted Kd. Although the data of the mineral-phase content are harder to obtain and may affected by the extraction process, the fitted kf and kb are independent of the value of Kd.
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TOP
ABSTRACT
INTRODUCTION
