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TECHNICAL REPORTS
Heavy Metals in the Environment

Heavy Metal Release from Contaminated Soils

Comparison of Column Leaching and Batch Extraction Results

Institute of Terrestrial Ecology, Swiss Federal Institute of Technology, Grabenstrasse 3, CH-8952 Schlieren, Switzerland

^* Corresponding author (kretzschmar{at}ito.umnw.ethz.ch)

Received for publication January 24, 2002.

	ABSTRACT

TOP ABSTRACT INTRODUCTION MATERIALS AND METHODS RESULTS AND DISCUSSION CONCLUSIONS REFERENCES

 
Heavy metals in soils may adversely affect environmental quality. In this study, we investigated the release of Zn, Cd, Pb, and Cu from four contaminated soils by column leaching and single and sequential batch extractions. Homogeneously packed soil columns were leached with 67 mL/g 10^-2 M CaCl₂ to investigate the exchangeable metal pool and subsequently with 1400 mL/g 10^-2 M CaCl₂ adjusted to pH 3 to study the potential of metal release in response to soil acidification. In two noncalcareous soils (pH 5.7 and 5.1), exchange by Ca resulted in pronounced release peaks for Zn and Cd that were coupled to the exchange of Mg by Ca, and 40 to 70% of total Zn and Cd contents were rapidly mobilized. These amounts compared well with exchangeable pools determined in single and sequential batch extractions. In two soils with near-neutral pH, the effluent concentrations of Zn and Cd were several orders of magnitude lower and no pronounced elution peaks were observed. This behavior was also observed for Cu and Pb in all four soils. When the soils were leached at pH 3, the column effluent patterns reflected the coupling of CaCO₃ dissolution (if present) and other proton buffering reactions, proton-induced metal release, and metal-specific readsorption within the soil column. Varying the flow rate by a factor of five had only minor effects on the release patterns. Overall, Ca exchange and subsequent acidification to pH 3 removed between 65 and 90% of total Zn, Cd, Pb, and Cu from the four contaminated soils.

Abbreviations: SSR, solution to soil ratio

	INTRODUCTION

TOP ABSTRACT INTRODUCTION MATERIALS AND METHODS RESULTS AND DISCUSSION CONCLUSIONS REFERENCES

 
CONTAMINATION OF SOILS with heavy metals is widespread and poses a long-term risk to ground water quality and ecosystem health. The environmental impact of heavy metal contaminants strongly depends on the metals speciation, mobility, and bioavailability in soil. Thus, risk assessments and soil remediation strategies must take these factors into account, in addition to the total metal concentrations. The chemical behavior of heavy metals in soils is controlled by a number of processes, including metal cation release from contamination source materials (e.g., fertilizer, sludge, smelter dust, ammunition, slag), cation exchange and specific adsorption onto surfaces of minerals and soil organic matter, and precipitation of secondary minerals (Manceau et al., 2000; McBride et al., 1997; McBride, 1999; Morin et al., 1999). The relative importance of these processes depends on soil composition and pH. In general, cation exchange reactions and complexation to organic matter are most important in acidic soils, while specific adsorption and precipitation become more important at near-neutral to alkaline pH values.

The dominating mineral species of heavy metals in contaminated soils can be identified by X-ray absorption fine structure (XAFS) spectroscopy, which provides information about the short-range coordination chemistry of the metal atoms (Hesterberg et al., 1997; Manceau et al., 2000; Scheinost et al., 2002). However, identification and quantification of the most mobile metal species with spectroscopic tools is still difficult. Therefore, several other approaches are commonly used to estimate so-called mobile, labile, or bioavailable pools of heavy metals in soil. The most important methods include: (i) single batch extraction of soil samples with salt solutions (e.g., NH₄NO₃ or CaCl₂), (ii) sequential batch extractions with increasingly harsh extractants designed to dissolve metals associated with different solid phases, (iii) isotope exchange methods, (iv) the diffusive gradients in thin films (DGT) technique, and (v) column leaching experiments. Single batch extractions form the basis of environmental regulations in many countries (Gupta et al., 1996; McLaughlin et al., 2000a,b). Sequential batch extractions are often used to obtain information about the possible binding forms of heavy metals in soil, although the fractions are strictly operationally defined and do not necessarily reflect true chemical speciation (Ahnstrom and Parker, 1999; Tessier et al., 1979; Zeien and Brümmer, 1989). Isotope exchange (Ahnstrom and Parker, 2001; Sinaj et al., 1999; Young et al., 2000) and DGT (Zhang et al., 1998, 2001) methods are often found to correlate best with plant uptake and are, therefore, useful to determine bioavailable pools. Finally, column leaching experiments provide information about heavy metal release and transport in soil, and are also useful for testing possible soil remediation or stabilization treatments, to assess heavy metal binding and desorption kinetics, or to study processes such as colloid-facilitated metal transport at the laboratory scale (Grolimund et al., 1996; Kedziorek et al., 1998; Kretzschmar and Sticher, 1997; Temminghoff et al., 1997).

Risk assessments and environmental regulations for polluted soils are often based on batch extractions of heavy metals, assuming that the results are related to the risk of metal leaching into ground water or plant uptake. However, a systematic comparison between mobile pools of heavy metals determined by single or sequential batch extractions and the amounts of metals released in column leaching experiments is currently lacking. Therefore, our objectives were to (i) investigate the mobilization of Cd, Zn, Pb, and Cu from several contaminated soils with different contamination histories in response to cation exchange and soil acidification, and (ii) compare the amounts of heavy metals mobilized in column leaching experiments with the results of conventional single and sequential batch extractions. Column effluent concentration patterns are discussed in the light of coupled chemical processes, such as cation exchange, adsorption, dissolution, and proton buffering.

	MATERIALS AND METHODS

TOP ABSTRACT INTRODUCTION MATERIALS AND METHODS RESULTS AND DISCUSSION CONCLUSIONS REFERENCES

 
Field Sites and Sampling
Four metal-contaminated soils were used in this study, which will be referred to as Soil 1 to Soil 4. Selected soil properties are summarized in Table 1. Soil 1 was collected in the East Midlands region of England (Carlton Forest Farm, Worksop) and is a well-drained, slightly acidic soil developed on Triassic sandstone. Until approximately 1980, large amounts of sewage sludge were applied to this soil, resulting in severe Zn pollution. Reduced plant growth due to Zn toxicity is still observed today. Soils 2 and 3 are located in northern France (Evin, Lille) and have developed from a quaternary clay deposit. Both soils are poorly drained, with a shallow ground water table approximately one meter deep. The major difference is that Soil 2 is forested with poplar (Populus spp.) trees, while Soil 3 is used for annual crop production and has a higher pH value due to liming (Table 1). Both soils are contaminated with Zn, Pb, and Cd by past atmospheric dust emissions from a large zinc–lead smelter, which started operation around 1884. In the late 1970s, filters were installed, reducing further soil contamination. Soil 4 is located in northern Switzerland (Mattenweg, Dornach) and has developed from a layer of weathered limestone gravel overlying a calcareous fluvial deposit. The soil is contaminated with Zn, Cd, and Cu by past dust emissions from a nearby brass smelter, which emitted metal-containing dusts over a period of about 80 years, before air filters were installed in the 1980s. Large soil samples were collected at all four sites from topsoil and subsoil horizons. The soils were air-dried and passed through a 2-mm plastic sieve to remove gravel and rocks.

A second set of experiments was conducted to study the further release of heavy metals induced by soil acidification after depletion of the Ca-exchangeable pools. For these experiments we used smaller columns containing only 1.5 g soil material, because we wanted to monitor the effluent over a much larger number of pore volumes (Table 2, Columns 2, 4, 6, and 10). The column conditioning with 10^-5 M CaCl₂ and leaching with 10^-2 M CaCl₂ solutions were performed as described above (Solutions A and B in Table 2), using the same leaching volumes per mass of soil. Subsequently, the columns were percolated with 1400 mL/g of 10^-2 M CaCl₂ solution adjusted to pH 3.0 by HCl addition (Solution C in Table 2). Column effluents were collected and analyzed for Al, Si, Fe, Mn, Ca, Mg, K, Zn, Cd, Pb, and Cu by inductively coupled plasma–atomic emission spectrometry (ICP–AES) (Liberty 200; Varian) equipped with an ultrasonic nebulizer (U-5000 AT; CETAC Technologies, Omaha, NE).

To study possible effects of release kinetics and metal readsorption on the observed release patterns during soil acidification, additional column experiments were performed with Soil 3 and 4. In two experiments labeled low Ca, the CaCl₂ concentration during the acid leaching step was reduced to 10^-5 M (Table 2, Columns 7 and 11). In two other experiments, labeled low flow, the flow rate during the acid leaching step was reduced to 0.05 mL/min (Table 2, Columns 8 and 12).

Column pore volumes and Péclet numbers reported in Table 2 were determined at the end of each experiment by injecting a short nitrate pulse (0.1 mL) and monitoring the nitrate breakthrough peak using on-line UV-Vis detection at a 220-nm wavelength. The flow rate was set to 0.25 mL/min except for Columns 8 and 12, where it was set to 0.05 mL/min. The average breakthrough time of the nitrate peak provides information about the column pore volume, while the width of the nitrate peak is related to dispersivity and can therefore be used to determine the column Péclet number (Villermaux, 1981). Note that the small columns filled with only 1.5 g of soil material had rather high Péclet numbers (Pe > 36), thus yielding satisfactory chromatographic resolution.

Batch Extractions
Several different single and sequential batch extraction methods were applied to all soils to compare the extractable metal pools with column leaching experiments. Single batch extractions were conducted using the following salt solutions and solution to soil ratios (SSR, in mL/g): 0.1 M NaNO₃ (SSR = 2.5) (Gupta and Aten, 1993), 0.1 M BaCl₂ (SSR = 30) (Hendershot and Duquette, 1986), 10 mM CaCl₂ (10 and 100, 24 h), and 1 M CaCl₂ (10 and 100, 24 h). A sequential selective extraction (SSE) procedure was performed consisting of seven extraction steps (Schwartz et al., 1999; Zeien and Brümmer, 1989). The hypothetical interpretation of the operational fractions F1 to F7 according to Zeien and Brümmer (1989) is given in parentheses: F1: 1 M NH₄NO₃, SSR = 25 (readily soluble and exchangeable); F2: 1 M NH₄–acetate, pH 6.0 (specifically adsorbed, CaCO₃ bound, and other weakly bound species); F3: 0.1 M NH₂OH–HCl plus 1 M NH₄–acetate, pH 6.0 (bound to Mn oxides); F4: 0.025 M NH₄EDTA, pH 4.6 (bound to organic substances); F5: 0.2 M NH₄–oxalate, pH 3.25 (bound to amorphous and poorly crystalline Fe oxides); F6: 0.1 M ascorbic acid in 0.2 M NH₄–oxalate, pH 3.25, in boiling water (bound to crystalline Fe oxides); and F7: HF–HNO₃–HCl (residual fraction). Zinc, Cd, Pb, and Cu in extracts were measured with flame atomic absorption spectrometry.

	RESULTS AND DISCUSSION

TOP ABSTRACT INTRODUCTION MATERIALS AND METHODS RESULTS AND DISCUSSION CONCLUSIONS REFERENCES

 
Metal Release by Calcium Exchange
Metal concentrations and pH in the column effluent of Soil 1 (Column 1, Table 2) in response to increased Ca concentration in the influent are presented in Fig. 1a . During the initial conditioning phase, only small amounts of Cd and Zn were released, possibly in association with mobile colloidal particles resulting from column packing. Increasing the influent Ca concentration to 10^-2 M resulted in a sharp and unretarded increase in effluent concentrations of Zn, Cd, and Mg. Correspondingly, effluent pH decreased from approximately 6.2 to 5.7, which is close to the soil's pH value determined in 10^-2 M CaCl₂ (Table 1). Due to the low cation exchange capacity and small amounts of exchangeable Mg in this soil, the Mg concentration in the effluent exhibited only a short peak and then rapidly decreased again to very low values. Effluent Cd and Zn concentrations also increased immediately after switching the influent to 10^-2 M CaCl₂ solution. However, their peaks were followed by a pronounced tailing. Note that Cd and Zn exhibited very similar release patterns, except that dissolved Zn concentrations were about 2000 times higher than those of Cd, reflecting the different contamination levels (Table 1). The similar release patterns for Cd and Zn are in good agreement with the sorption and transport behavior of both metals in acidic soils (Voegelin et al., 2001; Kretzschmar and Voegelin, 2001).

View larger version (22K): [in this window] [in a new window]   Fig. 1. Column effluent composition of Soil 1. (a) Preconditioning with 10-5 M CaCl2 and leaching of Ca-exchangeable metals with 10-2 M CaCl2 solution (pH 6). An effluent volume of 1 mL/g corresponds to approximately 2.6 pore volumes. (b) Leaching with 10-2 M CaCl2 adjusted to pH 3.0. An effluent volume of 1 mL/g corresponds to approximately 2.1 pore volumes. The first 100 mL/g corresponded to the same leaching sequence as in (a).

View larger version (23K): [in this window] [in a new window]   Fig. 2. Column effluent composition of Soil 2. (a) Preconditioning with 10-5 M CaCl2 and leaching of Ca-exchangeable metals with 10-2 M CaCl2 solution (pH 6). An effluent volume of 1 mL/g corresponds to approximately 1.7 pore volumes. (b) Leaching with 10-2 M CaCl2 adjusted to pH 3.0. An effluent volume of 1 mL/g corresponds to approximately 1.5 pore volumes. The first 100 mL/g corresponded to the same leaching sequence as in (a).

The effluent concentrations of Zn, Cd, Mg, and Ca from Soil 3 (Column 5, Table 2) are presented in Fig. 3a . Note that the concentrations of Cd and Zn are much lower than for Soils 1 and 2 and are, therefore, plotted on a µmol/L scale. Increasing Ca in the influent resulted only in slight increases in Cd and Zn concentrations in the effluent. The concentrations of both metals appeared to reach a plateau level, although the data are noisy due to analytical errors. In principle, the effluent patterns of both metals were similar to those of Pb in Soil 2 (Fig. 2a). The exchangeable pools of Cd and Zn were obviously not depleted at the end of the leaching period. Compared with the peak concentrations of Cd and Zn in Soil 2 (Fig. 2a), the concentrations in the effluent from Soil 3 were roughly 3000 times lower for Zn and 50 times lower for Cd, although the contamination level of both soils differed only by a factor of 3 to 4 (Table 1).

View larger version (24K): [in this window] [in a new window]   Fig. 3. Column effluent composition of Soil 3. (a) Preconditioning with 10-5 M CaCl2 and leaching of Ca-exchangeable metals with 10-2 M CaCl2 solution (pH 6). An effluent volume of 1 mL/g corresponds to approximately 1.9 pore volumes. (b) Leaching with 10-2 M CaCl2 adjusted to pH 3.0. An effluent volume of 1 mL/g corresponds to approximately 1.7 pore volumes. The first 100 mL/g corresponded to the same leaching sequence as in (a).

View larger version (25K): [in this window] [in a new window]   Fig. 4. Column effluent composition of Soil 4. (a) Preconditioning with 10-5 M CaCl2 and leaching of Ca-exchangeable metals with 10-2 M CaCl2 solution (pH 6). An effluent volume of 1 mL/g corresponds to approximately 1.6 pore volumes. (b) Leaching with 10-2 M CaCl2 adjusted to pH 3.0. An effluent volume of 1 mL/g corresponds to approximately 1.1 pore volumes. The first 100 mL/g corresponded to the same leaching sequence as in (a).

Metal Release by Acidification
The concentrations of Zn, Cd, Si, Al, and Mg in the effluent of Soil 1 in response to acidification are depicted in Fig. 1b. In the acidification experiments, the soils were first leached with the conditioning solution and with 10^-2 M CaCl₂ as discussed above, and then the influent was acidified to pH 3 (Table 2). Because Soil 1 is a noncalcareous, sandy soil with low acid neutralization capacity, the effluent pH decreased shortly after introducing the pH 3 solution. The release of Al was closely coupled to the pH front and also started at approximately 130 mL/g. In contrast, the concentrations of Zn, Cd, and Mg in the effluent increased immediately after switching to the acidified influent solution at 100 mL/g. While the peak effluent concentrations of Cd and Zn were approximately 10 times lower than in the precedent leaching with nonacidified CaCl₂ solution (Fig. 1a), metal release in response to acidification lasted much longer.

Similar metal elution patterns were also observed for Soil 2 in response to acidification (Fig. 2b). Again, Cd and Zn concentrations in the effluent increased immediately after switching to acidified influent. In contrast, the concentration increases of Al and Pb were retarded and appeared to be coupled to the pH front. Following the peak concentrations at about 270 and 230 mL/g, respectively, release of Al and Pb continued with long tailings over the entire duration of the experiment.

The coupling of Al and Pb release to pH breakthrough was even more obvious for Soil 3 (Fig. 3b), which has a higher acid neutralization capacity due the presence of 7 g/kg CaCO₃ (Table 1). The pH breakthrough was more strongly retarded and occurred at around 370 mL/g (effluent pH approximately 3.3), which is consistent with proton consumption of CaCO₃ dissolution and other buffering reactions. Effluent concentrations of Cd and Zn increased clearly before pH breakthrough occurred, although Zn was somewhat more retarded than Cd, probably due to stronger Zn adsorption to the soil column at pH of approximately 6.

Figure 4b shows the corresponding results for Soil 4, which has the highest acid neutralization capacity due to 66 g/kg CaCO₃ (Table 1). As a result, the pH breakthrough front was strongly retarded and the effluent pH only reached pH 3.3 at 1500 mL/g. Again, Al release occurred simultaneously with pH breakthrough, Cu release only slightly earlier than Al, and Cd and Zn release shortly after switching to the acidified influent. Compared with the other three soils, release peaks and the pH breakthrough front were much more diffuse, which is probably due to slow dissolution kinetics of carbonates and possibly other metal-bearing solid phases.

If the results for all four soils are considered together, a general trend in the elution patterns can be observed: a retarded acidification front results in clearly separated effluent peaks for Cd, Zn, Cu, Pb, and Al. The sequence of increasing retardation, Cd (-10.1) < Zn (-9.0) < Cu (-7.7) < Pb (-7.6) < Al (-5.0), reflects the sequence of increasing metal hydrolysis constants (log K, given in parentheses), which also correlate with cation adsorption affinity in soils (Alloway, 1995). At the acidification front, all metals (if present) are released from the soil material. However, mobilized Al and Pb, due to their higher adsorption affinity, readsorb in zones of higher pH within the column, and their appearance in the effluent is therefore coupled to the pH breakthrough front. To a slightly lesser extent, this also applies for Cu. In contrast, Cd and Zn have a lower affinity to the soil material and therefore travel faster than the retarded pH front. Thus, Cd and Zn concentrations in the effluent increase long before breakthrough of the acidification front. Since Zn is more strongly adsorbed at pH 6 to 7 than Cd, separation of the Cd and Zn effluent peaks can be observed in the soils (3 and 4) with near-neutral pH.

Readsorption and Kinetics
To demonstrate the effects of metal readsorption on the position of the metal release peaks, two additional experiments were conducted with Soils 3 and 4 using a lower Ca concentration (10^-5 M) for leaching with acidified solution. The results of these experiments (labeled low Ca) are depicted in Fig. 5a and 6a . In the experiment with Soil 3, effluent concentration peaks of Cd and Zn were now much more strongly retarded and occurred simultaneously with the acidification front (compare Fig. 3b and 5a). The effluent concentration peak of Pb was even more retarded and the peak concentration was much lower than at higher Ca concentration. Also, Al was released only at very low concentration levels. The effluent Ca concentration was initially about 0.6 mM due to CaCO₃ dissolution, and decreased to 10^-5 M after the CaCO₃ was depleted and pH breakthrough occurred. Very similar trends were also observed for Soil 4 (compare Fig. 4b and 6a). The effluent concentration peak of Cu was also more retarded than at higher Ca concentration and Cu concentrations increased simultaneously with the pH breakthrough front.

View larger version (26K): [in this window] [in a new window]   Fig. 5. Column effluent composition of Soil 3 during leaching with CaCl2 solutions adjusted to pH 3. (a) Experiment with a low Ca influent concentration of 10-5 M (an effluent volume of 1 mL/g corresponds to approximately 1.9 pore volumes). (b) Experiment at low flow rate of 0.05 mL/min (an effluent volume of 1 mL/g corresponds to approximately 1.7 pore volumes).

View larger version (27K): [in this window] [in a new window]   Fig. 6. Column effluent composition of Soil 4 during leaching with CaCl2 solutions adjusted to pH 3. (a) Experiment with a low Ca influent concentration of 10-5 M (an effluent volume of 1 mL/g corresponds to approximately 1.5 pore volumes). (b) Experiment at low flow rate of 0.05 mL/min (an effluent volume of 1 mL/g corresponds to approximately 1.2 pore volumes).

The interpretation of column leaching experiments may be further complicated by the occurrence of slow desorption or dissolution reactions, which could influence the metal release patterns. To assess the importance of kinetic effects, two experiments with Soils 3 and 4 were conducted at a five-times-lower flow rate (0.05 mL/min, labeled low flow). The resulting effluent pH and concentrations of Zn, Cd, and Cu or Pb are presented in Fig. 5b and 6b, respectively. Elution curves of Cd and Zn from Soil 3 were shifted to slightly larger effluent volumes (compare Fig. 3b and 5b). However, the effect was small compared with the Ca-concentration effect (Fig. 5a). The overall elution pattern was very similar at both flow rates. In case of Soil 4, the elution peaks of Cd and Zn were nearly unchanged at lower flow rate (compare Fig. 4b and 6b). The pH breakthrough was slightly more retarded when flow rate was decreased, which may be related to dissolution kinetics of CaCO₃. The position of the Cu elution peak was not affected by lowering the flow rate, but the peak became sharper, which is in agreement with the higher column Péclet number at lower flow rate (Table 2).

Comparison of Column and Batch Results
To compare column with batch results, we calculated the cumulative amounts of Cd, Zn, Pb, and Cu released from the four contaminated soils during leaching with CaCl₂ solution at pH of approximately 6. The results are summarized in Table 3. Comparison of the leached amounts with the total metal contents (Table 1) shows that Cd and Zn were most mobile, especially in the acidic Soils 1 and 2. The Ca-exchangeable amounts of Zn were even in the same range as those of major cations such as Mg and K (not shown). In comparison, much smaller fractions of the total Cd and Zn were released by Ca exchange from Soils 3 and 4.

Metal extraction in batch not only depends on the salt concentration and SSR, but also on the adsorption affinity of the exchanging cation. The smallest fractions of total Zn and Pb were mobilized with 0.1 M NaNO₃ at SSR 2.5 mL/g (Gupta and Aten, 1993), due to both the low adsorption affinity of monovalent Na⁺ and the low SSR. Somewhat more Zn and Pb were extracted with 0.1 M BaCl₂ at SSR 30 mL/g (Hendershot and Duquette, 1986) and with 1 M NH₄NO₃ at SSR 25 mL/g (Zeien and Brümmer, 1989), two extraction methods commonly used to determine exchangeable cations. With both methods, however, significantly less Zn and Pb were extracted than with 1 M CaCl₂. In the sequential extraction procedure of Zeien and Brümmer (1989), the extraction with 1 M NH₄NO₃ (F1, "exchangeable fraction") is followed by an extraction with 1 M NH₄–acetate (F2, "specifically adsorbed fraction"). The sum F1 + F2 compares well with the results of 1 M CaCl₂ batch extractions. While both F1 and F2 use NH₄⁺ as the exchanging cation, NO₃^- is used as anion in F1 and acetate in F2. Metal–acetate complexes are considerably more stable than respective metal–nitrate complexes. This leads to an effective decrease of the free metal cation activity in F2 and, consequently, to the release of more strongly adsorbed metals from the soil matrix. The decrease of metal activity by complexation with acetate in F2 has its analogy in the column leaching setup: in a flow-through column, the activity of the released metals in solution is kept low by constant leaching with fresh metal-free extractant solution, thereby favoring desorption reactions.

In Fig. 7 , we compare the amounts of Zn, Cd, Pb, and Cu released from soil columns in response to Ca exchange and acidification with the results of sequential selective extractions and 1 M CaCl₂ batch extractions. Several general conclusions can be drawn from this comparison:

1. For weakly adsorbed metals, such as Cd and Zn in acidic Soils 1 and 2, the amounts released in column leaching experiments by Ca exchange corresponded well with the amounts extracted in a single batch with 1 M CaCl₂ solution and with the sum of the first two fractions (F1 + F2) of the sequential selective extraction procedure.
   
2. For strongly adsorbing metals, such as Pb and Cu in all soils and Cd and Zn in Soils 3 and 4, the amounts released in columns by Ca exchange are always much lower than the amounts extracted in batch with 1 M CaCl₂ and the sum of F1 + F2.
   
3. In Soils 3 and 4, the sum F1 + F2 of the sequential batch extraction was generally higher than the amount extracted with 1 M CaCl₂ solution. This difference may be due to release of carbonate-bound metals by 1 M NH₄–acetate at pH 6 (F2), which could not be extracted with 1 M CaCl₂.
   
4. The observed metal release patterns in column experiments were related to the ratios of metals extracted with F1 and F2 of the sequential batch extraction. If F1 > F2, clear cation exchange patterns were observed in the effluent curves of column experiments in response to increased Ca concentration (e.g., Fig. 2a). If F1 < F2, the resulting elution curves exhibited a plateau at a low concentration (e.g., Fig. 4a) and the exchangeable pools were not depleted rapidly. If F2 is approximately equal to F1, a metal elution peak was followed by a strong tailing (e.g., Fig. 1a), which can be rationalized as initial metal release from sites with low metal affinity (corresponding to F1) followed by additional metal release from sites with higher metal affinity (corresponding to F2).
   

View larger version (58K): [in this window] [in a new window]   Fig. 7. Amounts of Cd, Zn, Pb, and Cu extracted in column leaching experiments, single batch extraction with 1 M CaCl2 solution at a solution to soil ratio (SSR) = 100 mL/g, and sequential batch extractions. All results are given in percent of the respective total metal concentrations (Table 1). The arrows on top of bars for column experiments indicate that the respective pools were not depleted at the end of the leaching period.

The cumulative amounts of metal cations released in column experiments by acidification are reported in Table 3 and they are also shown in Fig. 7 as percentage of total metal contents. This fraction is very poorly defined because it can include metal cations released from high-affinity sorption sites on organic matter and oxides, metals released by dissolution of metal-bearing solid phases, and even exchangeable metals in those cases where this pool was not depleted before acidification (Soils 3 and 4). Thus, the amounts of metals released by acidification only provide information about the potential of heavy metal release as a result of strong soil acidification. In general, between 70 and 90% of the total metal contents were mobilized by Ca exchange and subsequent leaching at pH 3.

	CONCLUSIONS

TOP ABSTRACT INTRODUCTION MATERIALS AND METHODS RESULTS AND DISCUSSION CONCLUSIONS REFERENCES

 
Carefully interpreted column leaching experiments provide valuable information on labile metal pools, metal sorption affinity, the coupling of simultaneous processes such as cation exchange, carbonate dissolution, and readsorption, as well as on the importance of slow reaction kinetics in metal leaching. In combination with results from spectroscopy and batch extractions, such information should prove useful for the calibration of transport models used to assess the risk of metal leaching at contaminated sites. Our results indicate that acidic inputs into contaminated soils may lead to increased Cd and Zn leaching long before the acid neutralization capacity of the soil is depleted. On the other hand, Pb and Cu leaching will be strongly retarded until the soil leachate itself becomes acidic.

	ACKNOWLEDGMENTS

 
We thank David G. Kinniburgh, Marc F. Benedetti, and Willem H. van Riemsdijk for many stimulating discussions during the course of the EU project FAMEST. Financial support of this research by the Swiss Ministry of Science and Education (Project BBW-Nr. 97-0116), in the framework of the EU project FAMEST (ENV4-CT97-0554), is gratefully acknowledged.

	REFERENCES

TOP ABSTRACT INTRODUCTION MATERIALS AND METHODS RESULTS AND DISCUSSION CONCLUSIONS REFERENCES

 

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