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DIVISION S-2—SOIL CHEMISTRY

Organic Ligand and pH Effects on Isotopically Exchangeable Cadmium in Polluted Soils

^a Dep. of Soil and Water, Adelaide University, PMB No. 1, Glen Osmond, South Australia 5064, Australia
^b CSIRO Land and Water, PMB No. 2, Glen Osmond, South Australia 5064, Australia
^c Laboratoire Sols et Environnement, ENSAIA-INRA/INPL, 2 avenue de la Forêt de Haye, BP 172, F-54505 Vandoeuvre-les-Nancy, France

* Corresponding author (collins{at}drecam.cea.fr)

	ABSTRACT

TOP ABSTRACT INTRODUCTION MATERIALS AND METHODS RESULTS AND DISCUSSION SUMMARY AND CONCLUSIONS REFERENCES

 
Organic ligands excreted by plant roots and pH changes in the rhizosphere are two factors that may modify the phytoavailability of Cd. The effect of these components on the quantity (isotopically exchangeable) and intensity (distribution between the solid and solution phases) of Cd in a polluted acidic and calcareous soil was examined in detail using a batch experimental system. The organic ligands included (0.25–5 mM) sodium tartrate, the free acid and sodium salt of citrate, histidine, and deoxymugineic acid (DMA). A reduction of pH was identified as the mechanism of Cd solubilization in the presence of some ligands. In other cases, however, organic ligands caused greater Cd solubilization than that because of pH changes alone. It was concluded that organic ligands might increase solution concentrations of Cd, in both soils, through a number of reaction mechanisms other than decreasing pH (e.g., varying the surface charge of the soil, cation exchange, aqueous metal complexation, etc.). In the acidic soil a reduction of pH from 5.7 to 3.8 only increased the quantity of isotopically exchangeable Cd (E value) by 17%, thus, indicating the existence of a recalcitrant nonlabile fraction of Cd in this soil. The organic ligands were also unable to significantly alter the E value of Cd. In contrast, the presence of organic ligands and decreases in pH elevated the E value of Cd in the calcareous soil by as much as 300%. It is postulated that conditions in the rhizosphere may not only increase the solution concentration of Cd, but also increase the quantity of phytoavailable Cd (L value).

Abbreviations: DMA, deoxymugineic acid • E, isotopically exchangeable Cd • ICP-AES, inductively coupled plasma atomic emission spectroscopy • K_d, distribution coefficients

	INTRODUCTION

TOP ABSTRACT INTRODUCTION MATERIALS AND METHODS RESULTS AND DISCUSSION SUMMARY AND CONCLUSIONS REFERENCES

 
PLANTS MAY MODIFY the physicochemical reactions of Cd in the soil environment (McLaughlin et al., 1998). Plant-induced changes of pH in the rhizosphere have been well documented and it is well known that pH has a critical influence on the distribution of Cd between the solid and solution phases of soils (Christensen, 1984; Tiller et al., 1984). Furthermore, as analytical techniques to measure plant-derived organic ligands in soils have improved, it is apparent that they are far more ubiquitous in the rhizosphere than originally assumed (Harter and Naidu, 1995; Jones, 1998).

Metal sorption to the solid phase of soils may be enhanced if complexation with an organic ligand results in an aqueous metal-ligand complex that has a high affinity for the soil solid phase (McBride, 1985). Similarly, the sorption of organic ligands may increase Cd sorption to soil surfaces by either increasing the negative electrostatic potential on the surfaces, which adsorb Cd, or by binding directly with the aqueous cation (Haas and Horowitz, 1986; Naidu and Harter, 1998). The presence of organic ligands may also inhibit the crystallization of some Fe (hydr)oxides. As the negative surface charge of these (hydr)oxides may be increased, the capacity for metal adsorption may also be elevated (Xue and Huang, 1995).

On the contrary, organic ligands may also stimulate the solubilization of Cd through a number of processes other than simple aqueous metal complexation reactions (Elliott and Denneny, 1982; Mench and Martin, 1991; Naidu and Harter, 1998). For example, sorption of positively charged organic ligands might increase solution concentrations of Cd by directly competing for cation-adsorption sites or by reducing the negative electrostatic potential on sorption surfaces (Levy and Francis, 1976; Chubin and Street, 1981). Organic ligands may also promote the desorption of Cd by dissolving minerals which adsorb the metal (Boyle and Voight, 1973; Jauregui and Reisenauer, 1982; Pohlman and McColl, 1986; Dinkelaker et al., 1989). Similarly, it is also possible that organic ligands may dissolve Cd precipitates in soils. As organic ligands may induce changes in soil pH (Dinkelaker et al., 1989), solubilization of solid bound Cd may take place because of decreases in pH. However, soil solution pH changes induced by plant-derived organic ligands are currently difficult to predict because it is largely unknown if they are excreted as the free acid or the alkali metal salt (Dinkelaker et al., 1989; Jones, 1998; Sakaguchi et al., 1999). As such, the former process may increase the solution concentration of Cd by decreasing the pH of the soil solution, whereas the latter may reduce concentrations of Cd in solution by inducing a rise in pH. In fact, distinctions between pH changes and other mechanisms of soil Cd solubilization in the presence of organic ligands have rarely been made (Naidu and Harter, 1998). Moreover, most studies have stressed metal adsorption and desorption reactions, which are surface processes, while precipitation and dissolution reactions may also be extremely important. This is particularly true for soils that are already polluted with Cd (Jopony and Young, 1994).

Isotopic dilution has been a useful technique to quantify the surface-adsorbed phase of an element that is in equilibrium with the solution phase (McAuliffe et al., 1947; Sumner and Bolt, 1962). In other words, a radioactive isotope, added to a preequilibrated soil suspension, will rapidly (hours to days) partition itself between the solid and solution phases in exactly the same manner as the surface-adsorbed stable element in equilibrium with the solution phase (McAuliffe et al., 1947). Therefore, the quantity of surface-adsorbed stable element, commonly referred to as the labile pool or E value, may be determined simply by measuring the distribution of the radioisotope between the solid and solution phases (K_d) and the concentration of stable element in solution (M):

[1]

Using the same principle, plants may be grown on isotopically equilibrated soils and the plant material used as the medium to measure concentrations of the radioactive and stable element—thereby replacing the soil solution used for determining E values. Although conceptually the same, results using plants as the sampling medium have been termed L values (Larsen, 1952).

It is these techniques which have provided evidence that the plant uptake of Cd in polluted soils may not always be limited to surface processes in the rhizosphere (Smolders et al., 1999; Gerard et al., 2000; Hutchinson et al., 2000). For example, Smolders et al. (1999) found that the labile pool of soil Cd available to Triticum aestivum L. (the L value) always exceeded the isotopically exchangeable pool of Cd measured in the laboratory (the E value) on 10 soils varying in pH (4.5–7.2) and Cd contamination. It was also observed that variations of pH across soils affected isotopically exchangeable Cd, but this phenomenon was not considered in depth. In addition, it has also consistently been demonstrated that high concentrations of EDTA (a synthetic organic ligand with strong metal-complexing abilities) increases the Cd E value of soils—presumably by the dissolution of Cd from occluded or interior sites in the soil (Nakhone and Young, 1993; Stanhope et al., 2000).

Therefore, the exudation of organic ligands and plant-induced changes of pH may arguably be two of the most important variables likely to affect the phytoavailability of Cd in the rhizosphere (Eriksson, 1989; Mench and Martin, 1991; Cieslinski et al., 1998). Despite this fact, extensive research identifying the forms of phytoavailable soil Cd and the factors controlling its quantity is still lacking in the literature. As this knowledge is essential to further develop sound and cost-effective remediation strategies, this study was conducted to determine the effects of H⁺ and plant-derived organic ligands on isotopically exchangeable Cd in two polluted soils varying in pH.

	MATERIALS AND METHODS

TOP ABSTRACT INTRODUCTION MATERIALS AND METHODS RESULTS AND DISCUSSION SUMMARY AND CONCLUSIONS REFERENCES

 
Soils
The A_p horizon of two soils (Hapludalfs), from a series of limed silt loam brown soils, was taken from Région Nord-Pas de Calais, France. The soils were located adjacent to two different Zn smelters and have been subject to Cd, Pb, and Zn pollution by the deposition of atmospheric particles. The first soil was called the ‘acidic soil’ while the other soil containing a proportion of calcium carbonate was termed the ‘calcareous soil’. A brief description of their physical and chemical properties is provided in Table 1. Before use, the soils were air-dried and sieved to <2 mm.

Determination of Distribution Coefficients and E Values of Cadmum
Distribution coefficients (K_d) and E values of Cd in the soils were measured by batch experiments using 60-mL polyethylene bottles. To these bottles a quantity of air-dry soil, equivalent to 2.5 g of oven-dry soil, was added with 25 mL of solution (e.g., organic ligand). All solutions were prepared with high purity water. Chloroform (50 µL) was also added at this time to minimize bacterial activity during isotopic equilibration.

A range of organic ligand concentrations from 0.25 to 5 mM were chosen and DMA was included with experiments using the calcareous soil. Variations of pH were facilitated by the addition of either 100 mM NaOH or HNO₃. The volume of these soil suspensions was made to 25 mL with high purity water.

Triplicate samples of each treatment were equilibrated by shaking on an end-over-end shaker at 40 rpm for 24 h. At this time 1 mL of carrier-free ¹⁰⁹Cd (0.24–9.41 kBq g^-1 soil) was added to the bottles and equilibrated for a further 24 h. To determine if these systems had reached equilibrium some bottles were shaken for an additional 48 h. No statistical difference, at the 0.05 probability level using the test of Newman-Keuls after ANOVA, was observed between the E value measured at these two equilibration periods using high purity water, 10 mM CaCl₂, 10 mM Ca(NO₃)₂, 0.5 mM solutions of all the organic ligands, and 5 mM solutions of citric acid and sodium citrate. Therefore, the shorter equilibration time was used throughout the experiments.

After equilibration the bottles were centrifuged at 1000 x g for 5 min to aid phase separation. The supernatant was filtered through cellulose nitrate filters (0.025 µm) into 20-mL polyethylene bottles and acidified to pH <2 with concentrated HNO₃ (Gerard et al., 2000). One milliliter was taken to measure the residual radioactivity of ¹⁰⁹Cd in solution while the remainder was refrigerated until analysis for stable Cd.

Protocol for Measuring Organic Ligand Sorption
Organic ligand sorption to the solid phase of the soils was determined by adapting the methodology of Jones and Brassington (1998) to the protocol used to determine the K_d and E values (i.e., 1:10 soil/solution ratio). After the addition of the organic ligand solutions and chloroform to the bottles containing the soils, they were shaken at 320 rpm for 10 min and then centrifuged at 1000 x g for 5 min to aid phase separation. Aliquots of <100 µL were taken from the supernatant and the concentration of organic ligand remaining in solution was analyzed by anion chromatography (described below). The amount of sorption was calculated by subtracting the quantity of organic ligand measured in solution from the initial concentration added. The concentration of organic ligand remaining in solution was also quantified after isotopic equilibration. The sorption of DMA was not measured in these experiments because of a lack of suitable instrumentation to accurately quantify this ligand. However, the pH of the soil suspensions was measured at all sampling times.

Analyses
The amount of ¹⁰⁹Cd added to the soil suspensions and that remaining in solution after equilibration was quantified by -spectrometry (Cobra 5003, Packard, Meriden, CT). Solution concentrations of stable Cd were analyzed by inductively coupled plasma atomic emission spectroscopy (ICP-AES) (JY 238, Jobin Yuon, Paris, France). As determined by the analyses of spiked samples and samples diluted with high purity water, the analytical error associated with ICP-AES measurements was <2% of the solution concentration of Cd. The detection limit of Cd using ICP-AES was limited to 0.5 µg L^-1. Solution concentrations of tartrate and citrate were measured using eluent-suppressed conductivity detection following NaOH gradient separation (1–35 mM) on a Dionex AS-11 anion-exchange analytical column equipped with a Dionex AG-11 guard column (Dionex Corp., Sunnyvale, CA). The purity of DMA from other organic acids and salts, after dialysis, was also established by the same methodology. Histidine was separated by the same procedure as that used for the organic acids but was quantified using direct-UV detection at 210 nm. Using this methodology the limit of detection for histidine was 10 µM.

Equations and Statistics
The K_d of Cd in the soils was calculated by the following equation:

[2]

where R is the quantity of carrier-free ¹⁰⁹Cd added (Bq L^-1), r is the quantity of ¹⁰⁹Cd remaining in solution after equilibration (Bq L^-1), and L/S equals the liquid/solid ratio (L kg^-1 soil)

The E value calculated from Eq. [1] only represents surface-adsorbed metal in equilibrium with the solution phase. Therefore, the total solution concentration of metal must be added to the E value of Eq. [1] to obtain the total isotopically exchangeable metal in the soil–solution system (Smolders et al., 1999):

[3]

Equation [3] was used in this study because it is considered that the total isotopically exchangeable metal best represents the quantity of metal available for plant uptake (Smolders et al., 1999).

Statistical differences between treatment means, after ANOVA, were calculated using the test of Newman-Keuls (at the 0.05 probability level) using the statistical computer program STATITCF, version 5 (ITCF, Biogneville, France). Nonlinear regressions were calculated using SIGMAPLOT, version 5.0 (SPSS Science, Chicago, IL).

	RESULTS AND DISCUSSION

TOP ABSTRACT INTRODUCTION MATERIALS AND METHODS RESULTS AND DISCUSSION SUMMARY AND CONCLUSIONS REFERENCES

 
Sorption of Organic Ligands
In general, there was little sorption of the organic ligands, except histidine, in both soils (Table 2). For most ligands, sorption was positively related to increases in the solution concentration of the ligand. However, as pH was not buffered in these experiments, varying the concentration of most ligands resulted in changes of pH (data not shown). Nevertheless, despite previous results indicating that pH may affect citric acid and sodium citrate sorption to acidic soils at concentrations >2 mM (Jones and Darrah, 1994), there were no significant differences between the sorption of these ligands to the soils used in these experiments. The sorption of tartrate was also unaffected by soil type. As sorption of these ligands was unaffected by changes in pH or soil type, it indicates that neither the pK_a of the ligand's functional groups, complexation with metals, nor the surface charge of the soils had any observable effect on their sorption behavior in these experiments.

Although the sorption of DMA was not measured, its sorption behavior should be similar to that observed for the negatively charged organic acids because its isoelectric point would be at a pH between pK₂ (pH 2.76) and pK₃ (pH 3.4) (Sugiura et al., 1981).

Organic Ligand Concentrations after Isotopic Equilibration
Preliminary results, as measured by plating techniques, indicated that the addition of 50 µL of chloroform to the soil suspensions only minimized microbial activity. Therefore, to accurately assess the influence of organic ligands on the chemistry of Cd in these soils, the solution concentration of the ligand after isotopic equilibration was also measured.

In the acidic soil, significant decreases in the solution concentration of sodium citrate and tartrate were observed at 0.25 mM and for citric acid from 0.25 to 1.0 mM. At the conclusion of the equilibration period histidine was also not detected for all treatments in the acidic soil. In the calcareous soil, significant decreases in the concentration of citric acid and sodium citrate at 0.25 mM and histidine at all concentrations were observed.

After equilibration it was noted that the pH of these solutions had decreased by as much as one pH unit (data not shown). Therefore, decreases in pH may have (i) increased the quantity of positively charged histidine molecules and, thus, enhanced histidine sorption, or (ii) increased the positive surface charge of the soils enhancing sorption of negatively charged metal-ligand complexes or the ligands themselves. A reduction in the solution concentration of these ligands may also have been simply the result of biodegradation.

Organic ligands other than citrate, tartrate, and histidine were not detected in the solutions. Therefore, changes in the K_d or E value of Cd in the presence of the added ligands was not due, in part, to microbially produced organic ligands. Nevertheless, it was assumed that further losses of organic ligands from solution, after the sorption experiments (>10 min), were a result of biodegradation. Hence, discussion on the effects of organic ligands on the behavior of Cd in these soils is based on the concentration of ligand remaining in solution, after isotopic equilibration, and the quantity of ligand, which was rapidly sorbed to the soil solid phase. This assumption may add a source of minor uncertainty to the conclusions of Cd solubilization mechanisms in the presence of some of these ligands.

The Distribution Coefficients of Cadmium as Influenced by pH
The K_d values of Cd in the soils were strongly pH-dependent (Fig. 1) , confirming previous observations on the distribution of Cd between the solid and solution phases of soils (Christensen, 1984; Tiller et al., 1984; Elliott et al., 1986; Filius et al., 1998). Over the pH ranges examined in these soils, the partitioning of Cd between the solid and solution phases was characterized by sigmoidal shaped curves. At lower pH values these curves were distinguished by steep changes in the quantity of Cd associated with the solid phase. The natural pH of the acidic soil after isotopic equilibration (pH 5.7) was located within this pH range, whereas the natural pH of the calcareous soil (pH 7.5) represented the end point of this range.

View larger version (10K): [in this window] [in a new window]   Fig. 1. The distribution of Cd between the solid and solution phases (represented as Kd) for the soils as affected by changes of pH. Values and error bars represent the mean and standard deviation of triplicate measurements. The five-parameter nonlinear regression fitted for the acidic soil was Kd = 48.4 + 2499/(1 + exp[-(pH-5.77)/0.18]0.86: R2 = 0.99, P < 0.0001 and for the calcareous soil was Kd = 112.4 + 3421/(1 + exp[-(pH-7.49)/0.0034]0.0112: R2 = 1.00, P < 0.0001. See text for explanation of equations.

Effect of Organic Ligands on the Distribution Coefficients of Cadmium in the Acidic Soil
The type and concentration of organic ligand affected the distribution of Cd between the solid and solution phases of the acidic soil (Fig. 2) . To distinguish between pH and other processes (e.g., chelation) affecting K_d, the data were compared with those predicted at the same pH value in the absence of the ligand. The values of K_d (y) over the pH range examined in this soil were obtained by fitting the data presented in Fig. 1 by a five-parameter nonlinear regression:

[4]

where y0 is the minimum K_d value, a is (maximum K_d - minimum K_d), x is pH, x0 is pH₅₀, b is a constant describing the linearity of the regression, and c is a measure indicating the sigmoidal quality of the regression.

View larger version (32K): [in this window] [in a new window]   Fig. 2. The Kd of Cd in the acidic soil as affected by organic ligands compared with estimated values at the same equilibrium pH in the absence of ligand. Values and error bars of Kd in the presence of ligands represent the mean and standard deviation of triplicate measurements. The Kd values in the absence of ligands were calculated using the five-parameter nonlinear regression derived from the data of Fig. 1 and the mean pH values measured from triplicate samples containing the ligands.

Although a general increase in the solution concentration of Cd occurred as concentrations of citric acid and tartrate increased, the proportion accounted for by pH also increased. In conjunction, increasing concentrations of these ligands resulted in a lowering of the equilibrium pH of the solution. Therefore, if increasing ligand concentrations are associated with decreases in pH then the K_d of Cd becomes increasingly controlled by the concentration of H⁺ in solution. These observations are consistent with the fact that the stability of Cd-ligand complexes decrease with a reduction in pH.

At higher pH values, where the K_d was lower than that predicted by pH alone, there are a number of processes that could have been responsible for increased Cd desorption and dissolution in the presence of sodium citrate and tartrate. Calculations made with the computer program GEOCHEM-PC, version 2.0 (Parker et al., 1995) indicated that the metal-ligand complexes of citrate and tartrate either carried a net zero or negative charge in the pH range examined in these experiments. Similar calculations predicted that the ligands themselves had a net negative charge. Therefore, elevated solution concentrations of Cd in the presence of the sodium salts of citrate and tartrate were not a result of a process similar to that observed for histidine. It is also highly unlikely that Na⁺ from the organic ligand solutions was contributing to Cd desorption based on the results of Naidu and Harter (1998) and Naidu et al. (1994). Therefore, sodium citrate and tartrate were increasing the solution concentration of Cd by forming nonsorbing soluble complexes with Cd thereby reducing Cd²⁺ (and possibly CdOH⁺) activities in the soil solution which would lead to the desorption of solid bound exchangeable Cd or the ligands were dissolving Cd precipitates or adsorption surfaces (e.g., Fe [hydr]oxides).

Effect of Organic Ligands on the Distribution Coefficients of Cadmium in the Calcareous Soil
The presence of organic ligands also decreased the K_d of Cd in the calcareous soil (Fig. 3) . A sigmoidal regression was also applied to the data in Fig. 1 to distinguish between pH and other possible mechanisms that may have increased the solution concentration of Cd in this soil. Increasing concentrations of citric acid decreased the pH of the soil solutions and the K_d's of Cd were not significantly different to those values calculated at the same pH in the absence of the ligand. This observation is similar to the results of the experiments with the acidic soil where decreases of pH generated metal-ligand instability and Cd desorption and dissolution was controlled by the solution concentration of H⁺.

View larger version (38K): [in this window] [in a new window]   Fig. 3. The Kd of Cd in the calcareous soil as affected by organic ligands compared with estimated values at the same equilibrium pH in the absence of ligand. Values and error bars of Kd in the presence of ligands represent the mean and standard deviation of triplicate measurements. The Kd values in the absence of ligands were calculated using the five-parameter nonlinear regression derived from the data of Fig. 1 and the mean pH values measured from triplicate samples containing the ligands.

Further evidence that chelation was affecting the solubilization of Cd and not the dissolution of adsorbing surfaces, was found with increases in the concentration of Fe following a slightly different order to that observed for Cd: DMA > citrate > histidine tartrate (data not shown). If the dissolution of adsorbing surfaces, such as Fe (hydr)oxides, was primarily responsible for the solubilization of Cd, then increases in the solution concentration of Cd should be expected to reflect that observed for increases in solution concentrations of Fe.

Cadmium E Values as Affected by pH and Organic Ligands
Despite the large effect that changes of pH had on the distribution of Cd between the solid and solution phases in the acidic soil, it had little consequence on the E value. For example, only at pH 3.8 was the E value significantly higher than that measured at pH values > 5.9. Furthermore, the presence of organic ligands did not significantly alter the E value of Cd in this soil when compared with values calculated at the same pH in the absence of the ligand (Fig. 4) . Further experiments conducted with 20 mM EDTA also indicated that the E value was indifferent to high concentrations of organic ligands with strong metal-complexing characteristics (E = 15.1 mg kg^-1). These results indicate that the E value of this soil is a stable pool of metal despite it only representing 65% of the total soil Cd. For example, when the equilibrium pH was reduced to 3.8 this percentage only increased to 75%, suggesting that acid digestion may be necessary to liberate the remaining nonisotopically exchangeable Cd from interior or occluded sites. As such, under relevant environmental conditions, a part of the Cd in this soil remains nonisotopically exchangeable and, furthermore, does not readily replenish that portion which is isotopically exchangeable.

View larger version (8K): [in this window] [in a new window]   Fig. 4. The E value of Cd in the acidic soil as affected by pH and the presence of organic ligands. The fitted line is a linear regression of the pH data where E = -0.8683pH + 18.223: R2 = 0.91, P > 0.001. The symbol () and error bars represent the mean value and standard deviation of triplicate measurements of pH. The other symbols represent: (0.25–5 mM) citric acid (x), (0.25–5 mM) sodium citrate (o), (0.25–1 mM) histidine (), and (0.25–1 mM) tartrate ().

View larger version (11K): [in this window] [in a new window]   Fig. 5. The E value of Cd in the calcareous soil as affected by pH and the presence of citric acid. The solid line represents the fitted five-parameter non-linear regression of E = 3.44 + 7.83/(1 + exp[-(pH-8.23)/-0.41]21: R2 = 0.99, P < 0.0001. Hashed lines represent 95% confidence limits. The symbol () represents the mean E value of Cd, in the absence of ligands, measured with triplicate samples. The symbol (x) and error bars represent the mean value and standard deviation of triplicate measurements of (0.25–5 mM) concentrations of citric acid.

View larger version (8K): [in this window] [in a new window]   Fig. 6. The effect of organic ligands on the E value of Cd in the calcareous soil. Values represent the ratio of the E value measured in the presence of the ligand against that measured in the absence of the ligand at pH 7.5. Symbols and error bars represent the mean and standard deviation of triplicate measurements. The symbols represent: (0.25–5 mM) sodium citrate (), (0.25–1 mM) histidine (), (0.5–1 mM) tartrate (), and (0.25–0.5 mM) DMA (+).

	SUMMARY AND CONCLUSIONS

TOP ABSTRACT INTRODUCTION MATERIALS AND METHODS RESULTS AND DISCUSSION SUMMARY AND CONCLUSIONS REFERENCES

 
In the present experiments, the E value of the acidic soil measured in water represented a major portion (if not all) of the Cd that maybe potentially phytoavailable. This conclusion was supported by the fact that decreases of pH and variations of solution composition, although greatly effecting K_d, had no significant effect on the size of the isotopically exchangeable pool of Cd. Based on these observations it is unlikely that conditions in the rhizosphere would alter the isotopic exchange properties of Cd in this soil. Therefore, from a practical point of view, the E value, regardless of the solution composition used for isotopic equilibration, is a good indication of the total quantity of Cd that would be phytoavailable (L value) in this soil. In contrast, variations of the E value in the calcareous soil indicated that isotopically exchangeable Cd measured in water may not be an accurate representation of all the metal potentially available in the presence of root-exuded organic ligands or upon acidification.

This study highlights the fact that the E value will only be a reliable indication of the quantity of phytoavailable Cd (L value) under one condition—when the isotopically exchangeable Cd is insensitive to changes of K_d. This condition was only met on the acidic soil. However, in soils where isotopically exchangeable Cd varies with K_d, it will also be possible, but much more difficult, to predict the quantity of phytoavailable Cd, but only if the K_d of Cd measured in the laboratory determined E values is exactly the same as the K_d in the rhizosphere. As such, in these types of soils, plant-induced decreases of pH and the exudation of organic ligands will increase the quantity of isotopically exchangeable Cd and, therefore, may account for previously observed differences between L and E values of Cd (Smolders et al., 1999; Gerard et al., 2000; Hutchinson et al., 2000).

Received for publication October 9, 2001.

	REFERENCES

TOP ABSTRACT INTRODUCTION MATERIALS AND METHODS RESULTS AND DISCUSSION SUMMARY AND CONCLUSIONS REFERENCES

 

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