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Published online 9 August 2005
Published in J Environ Qual 34:1530-1538 (2005)
DOI: 10.2134/jeq2004.0385
© 2005 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

Cadmium Solubility and Sorption in a Long-Term Sludge-Amended Arable Soil

Petra Bergkvist, Dan Berggren and Nicholas Jarvis*

Department of Soil Sciences, SLU, Box 7014, 750 07 Uppsala, Sweden

* Corresponding author (nicholas.jarvis{at}mv.slu.se)

Received for publication October 14, 2004.

   ABSTRACT
TOP
 ABSTRACT
INTRODUCTION
 MATERIALS AND METHODS
 RESULTS AND DISCUSSION
 CONCLUSIONS
 REFERENCES
 
Cadmium solubility and sorption in an arable clay loam soil that had received sewage sludge for 41 years were compared to an unsludged control in batch studies. Soil pH dominated Cd sorption, explaining >92% of the variation in Kd values in both treatments. At any pH, Cd sorption was apparently slightly but significantly (p < 0.05) smaller in the sludge-amended soil compared to the control, even though the organic carbon content was 70% larger and the ammonium oxalate–extractable iron content was roughly doubled. Correction for dissolved organic carbon (DOC) complexation with the speciation model WHAM reduced the difference in sorption between treatments, but the sludged soil still had significantly smaller Kd values (p < 0.01). Batch equilibrations without addition of Cd showed that there was no significant difference in the solubility of "native" cadmium (defined as EDTA-extractable Cd) in sludged and control soils. The reason for the lack of increase in Cd sorption in the sludge-amended soil has not been established, but it may be due to competition for sorption sites on humic compounds with sludge-derived Fe and trace metals such as zinc. The fact that the pyrophosphate-extractable (i.e., organically associated) iron content was seven times larger in the sludged soil provides some supporting evidence for this hypothesis.

Abbreviations: CEC, cation exchange capacity • DOC, dissolved organic carbon • SS, sewage sludge


   INTRODUCTION
TOP
 ABSTRACT
 INTRODUCTION
MATERIALS AND METHODS
 RESULTS AND DISCUSSION
 CONCLUSIONS
 REFERENCES
 
THE USE OF SEWAGE SLUDGE as a fertilizer in arable soil is a subject of concern and dispute. As a source of plant nutrients (especially phosphorus and nitrogen) and organic matter, sewage sludge is a beneficial soil amendment, especially for arable soils low in organic matter. However, as sewage sludge is a conglomerate of societal wastes, increased concentrations of toxic substances, including trace metals, are generally found in sludge-amended soils. Cadmium is recognized as being one of the most mobile trace metals circulating in the environment and can readily enter the human food chain, being more weakly bound to soil constituents compared to many other trace metals (Brümmer et al., 1986). Cadmium adsorption in soil is strongly controlled by soil pH (Christensen, 1984; McBride et al., 1997), and is also influenced by a range of soil constituents, including clay minerals (Zachara et al., 1992) and Mn, Al, and Fe oxides and hydroxides (Benjamin and Leckie, 1981; Fu et al., 1991; Bolton and Evans, 1996). Soil organic matter is usually identified as the soil constituent with the largest influence on Cd sorption (e.g., Christensen, 1989; Lee et al., 1996; Gray et al., 1999) and land application of sludge will increase soil organic matter contents above background levels. Compared to native soil organic matter, sludge-derived humic compounds seem to be more recalcitrant (Terry et al., 1979; Witter, 1996) and also have smaller contents of carboxyls and phenolic hydroxyls, the functional groups that are important for metal sorption (Sposito et al., 1982; Boyd and Sommers, 1990). Humic compounds are also subject to dissociation, releasing dissolved organic matter into the soil solution (Reemtsma et al., 1999). These compounds possess a high affinity for Cd and other trace metals, leading to a potentially enhanced solubility (Christensen, 1985; Neal and Sposito, 1986).

Besides trace metals, sewage sludge also contains adsorptive organic and inorganic components, which may influence the solubility of the added metals in sludge-amended soils in a long-term perspective. Legislative standards for sludge use in the United States partly rely on the sludge "protection" hypothesis (Chaney and Ryan, 1993), which suggests that trace metal sorption will be permanently enhanced due to a high sorption affinity in the inorganic fractions of sludge (i.e., carbonates, phosphates, and amorphous Mn and Fe oxides). Sorption experiments conducted on sludge-amended soils after removal of organic matter and/or Mn and Fe oxides suggest that these inorganic fractions can contribute significantly to Cd sorption (Li et al., 2001; Hettiarachchi et al., 2003). However, other batch studies have failed to show any persistent increase in trace metal sorption in sludged soils (O'Connor et al., 1983; Cline and O'Connor, 1984; Hooda and Alloway, 1994), while the evidence for sludge protection from field experiments may not always be unequivocal. For example, the observation that trace metal concentrations in plant tissues "plateau" while total soil metal contents are still increasing, has been attributed to sludge protection, but it can also be explained by plant physiological mechanisms (Hamon et al., 1999).

Thus, the long-term fate of sludge-borne Cd is still not completely understood. The aim of this study was to investigate the influence of long-term (41 yr) sewage sludge additions on Cd sorption and solubility in batch experiments performed on samples taken from sludge-amended and control treatments in a clay loam soil. The speciation model WHAM (Tipping, 1994) was used to investigate the impacts of measured DOC concentrations on Cd sorption, and the results of the batch experiments are discussed in relation to soil pH and changes in soil organic matter, Fe, Al, and Mn contents resulting from the sludge applications.


   MATERIALS AND METHODS
TOP
 ABSTRACT
 INTRODUCTION
 MATERIALS AND METHODS
RESULTS AND DISCUSSION
 CONCLUSIONS
 REFERENCES
 
Site Description and Soil Sampling
The experimental field from which the soil samples were collected is situated at the Swedish University of Agricultural Sciences, Ultuna, located about 8 km south of Uppsala (59°48' N, 17°39' E). According to FAO (1998) the soil is a Eutric Cambisol and a Typic Eutrocrept according to Soil Taxonomy (Kirchmann et al., 1996). The soil is a slightly calcareous clay loam. The experiment is arranged as a randomized block design, with four replicate plots for each treatment. The plots are 2 x 2 m in size, separated by wooden frames, extending 10 cm above the soil surface and sunk approximately 30 cm into the soil. Each plot has been hand-treated for all fieldwork since 1956 (Kirchmann et al., 1994). In the period 1956–1997, the plots were cultivated with annual crops, mostly spring-sown cereals and rape. The experiment includes 15 treatments, all continuously supplied with inorganic or organic N fertilizers since 1956. In this study, two of the treatments were selected for further investigation: sewage sludge amended soil (SS treatment) which biennially receives anaerobically digested and dewatered sewage sludge originating from the Uppsala wastewater treatment plant, and calcium nitrate fertilized soil (control). During 1956–1997, 25 kg dry matter m–2 (0.25 Mg ha–1) of sewage sludge had been supplied to the SS treatment plots, containing approximately 7.2 kg organic carbon m–2 (Kirchmann et al., 1994). No additional N fertilizer was supplied to the SS treatment, while the control received annually 80 kg N as calcium nitrate (15.5% N). Both treatments received P and K as combined fertilizer nearly every year. The inorganic fertilizers were added in spring, while sewage sludge was supplied in the autumn, followed by spade tillage to approximately 20 cm depth.

The total amount of Cd from all sources (sewage sludge, P fertilizer, and atmospheric deposition) supplied between 1956 and 1997 was estimated at approximately 153 mg Cd m–2 in the SS treatment and 13 mg m–2 in the control (Bergkvist et al., 2003). The Cd loading rate to the SS plots was therefore approximately four times less than the EU limit for sludge allowed under Directive 86/278/EEC150 (15 mg Cd m–2 yr–1), but greatly exceeded that allowed under Swedish legislation (0.075 mg m–2 yr–1). Mass balance calculations showed that approximately 92% of the cadmium added in sludge remained in the topsoil (Bergkvist et al., 2003).

Until 1975, a combination of Al and Fe salts was added to precipitate phosphorus in the Uppsala sewage treatment plant, but thereafter only Fe salts (FeCl2) were used. Between 1956 and 1997, the total supply of sludge-borne Fe, Al, and Mn was estimated at approximately 1.5, 0.2, and 0.06 kg m–2, respectively, based on monthly measurements reported since 1968 (Mn), 1972 (Fe), and 1975 (Al) (Ernst-Olof Swedling, Uppsala municipality, personal communication).

Initially, soil was sampled in 1997 from all four plots in each treatment at depths of 0 to 20, 20 to 25, and 25 to 30 cm to identify changes in topsoil thicknesses since the start of the experiment in 1956 and to measure important soil properties influencing cadmium adsorption and solubility (Bergkvist et al., 2003). The topsoil layer was calculated to have increased from 20 cm in 1956 to 28 cm in the SS treatment and 23 cm in the control (Bergkvist et al., 2003), caused partly by soil recovery as soil compaction by field machinery had ceased and partly by the repeated additions of sewage sludge. As the elevation of the soil surface was different between the treatments, the base of the topsoil was considered as the reference level for further comparisons. A second sampling was then performed to obtain soil for the batch experiments from the entire topsoil layer (i.e., 0–28/0–23 cm), a 7-cm-thick soil layer below the topsoil and, in the control treatment, two additional subsoil layers each 10 cm thick (i.e., at 30–40 and 40–50 cm depth). Samples were taken from each plot by a soil auger and were then pooled (Replicates 1 + 2 and Replicates 3 + 4), yielding two main replicates of each layer in the profile for each treatment.

Soil Analyses
All analyses were performed on air-dried soil samples with an aggregate size of <2 mm. Soil particle size distribution and soil pH in 0.01 M CaCl2 were measured using standard methods. Potential cation exchange capacity was calculated as the sum of exchangeable base cations, and titratable acidity at pH 7.8. Calcium, sodium, magnesium, and potassium were analyzed on a PerkinElmer (Wellesley, MA) AA300 AAS. Total cadmium was measured on 7 M HNO3 extracts. A measure of plant available Cd (Roca and Pomares, 1991) was obtained by extraction in 0.025 M Na2EDTA solution (Bergkvist et al., 2003). Extracts were filtered through Schleicher & Schuell (Dassel, Germany) membrane filters to retain particles greater than 0.2 µm, and then analyzed using a PerkinElmer 4110 ZL Zeeman AAS. The carbon content in soil samples was measured in a LECO (St. Joseph, MI) CHN 932 analyzer. This is considered equivalent to the organic carbon content, since carbonate C was only found deeper than 50 cm in the profile.

Ammonium oxalate (0.2 M, pH 3 in the dark) was used to extract crystalline and noncrystalline forms of Al (McKeague et al., 1971) and noncrystalline iron oxides (McKeague and Day, 1966). Ammonium pyrophosphate (0.1 M, pH 10) was used to extract Al and Fe associated with organic matter, either as complexes or as amorphous (oxy)hydroxides (McKeague and Schuppli, 1982; McKeague, 1967). The difference between NH4–oxalate and pyrophosphate-extractable Al was regarded as inorganic Al phases (Wang et al., 1986). The extraction procedures followed the methods described by Karltun et al. (2000). Easily reducible manganese (Mn) oxides and hydroxides were extracted in acidic 0.1 M hydroxylamine hydrochloride (Chao, 1972), according to standard procedures (Gambrell, 1996). Iron and aluminium extracted by ammonium oxalate as well as manganese were analyzed on a Jobin Yvon (Edison, NJ) JY24 ICP. Pyrophosphate-extracted Al and Fe were analyzed by plasma emission spectroscopy using a PerkinElmer AAnalyst 300 AAS.

Batch Experiments
Two series of batch experiments were performed to investigate Cd sorption and solubility. The first series of unreplicated equilibrations was performed in 1999 on one of the pooled samples from sludge-amended and control soils at three selected pH values (including the in situ value) within the range 4.6 to 6.6. This experiment was not replicated, since the within-treatment variability in Cd sorption was not expected to be significant. To check this, a second series of replicated equilibrations was performed in 2003 on both pooled samples from the topsoils of both treatments, this time at two controlled pH values (5.5 and 6.5).

At each selected pH value, three initial Cd concentrations were obtained by the addition of 100 µL Cd(NO3)2 solutions up to a maximum of 0.375 mg Cd L–1, including a zero Cd addition treatment to obtain the in situ equilibrium concentration. This yielded equilibrium Cd solution concentrations in a low concentration range (0.14–35.70 µg L–1) relevant for sludge-amended soils in a legislative context. In each batch experiment, 2 g air-dried soil was equilibrated by end-over-end rotation during 24 h at 10°C in a matrix solution of 20 mL 0.01 M Ca(NO3)2 solution, followed by centrifugation at 1000 rpm for 5 min. An equilibrium time of 24 h was chosen as Christensen (1984) reported that Cd sorption at low concentrations was a fast process reaching a constant distribution between soil and solution within 1 h. The dissolved organic carbon concentration has also been reported to reach equilibrium within 24 h in batch studies (You et al., 1999). The temperature was selected to be close to the yearly average soil temperature in the district (5.7°C) and the matrix solution was chosen to mimic the soil solution composition and ionic strength in arable soil (Filius et al., 1998; Holm et al., 1998). The pH adjustments were made by additions of 100 µL HNO3 or NaOH. After equilibration, the centrifuged solutions were divided into two vessels. The pH was measured on unfiltered solution, while the solutions in the other vessel were filtered through Pall (East Hills, NY) Acrodisc syringe filters retaining particles greater than 0.2 µm. The filtered solutions were acidified before measuring Cd concentrations on a PerkinElmer 4110 ZL Zeeman atomic absorption spectrophotometer.

The DOC concentrations were measured at equilibrium at each pH value, at zero and maximum Cd additions. The solutions were filtered through Pall Acrodisc syringe filters (0.2 µm) before measuring DOC with a Shimadzu (Kyoto, Japan) 5000A TOC analyzer.

Modeling Cadmium Speciation in WHAM
The solution speciation of Cd was calculated using Version 5 of the chemical equilibration model for aqueous phases, WHAM-W (Tipping, 1994), to investigate the effects of the different DOC concentrations in sludged and control soils on cadmium sorption and solubility measured in the batch experiments. WHAM comprises submodels for specific binding by humic substances, nonspecific binding by counter ion accumulation, and inorganic solution chemistry. In the calculations, we assumed that (i) organic ligands present behaved as fulvic acids, since these are the dominant water-soluble humic compounds, (ii) the C content of dissolved humic compounds was 50% (Tipping, 2002), and (iii) inorganic ligands such as Cl–, SO42–, and OH– were present in concentrations that would not significantly influence Cd speciation. The model input consisted of the known temperature and the pH and Cd and DOC concentrations measured in the 0.01 M Ca(NO3)2 extracts without Cd addition. Calcium is the dominating cation in our soil, occupying 92 to 95% of soil cation exchange capacity (CEC), thus only the occurrence of the matrix concentration of 0.01 M Ca and 0.02 M NO3– was accounted for in the model input, besides the measured Cd and DOC concentrations.

Calculation of Distribution Coefficients
The amount of Cd adsorbed or desorbed, {Delta}S (µg kg–1), in the batch equilibrations was calculated as:

[1]
where Ca is the added Cd concentration (µg L–1), Ceq is the equilibrium Cd concentration in solution (µg L–1), and R is the solution to soil ratio (10 L kg–1). The limited range of low concentrations of Cd used in our study resulted in a linear partitioning between sorbed and equilibrium solution Cd concentrations. Therefore, at each pH, "apparent" sorption distribution coefficients (Kd-app) were calculated by linear regression of {Delta}S against Ceq (r2 > 0.97 in all cases).

The adsorption distribution coefficient for Cd between soil and Cd2+ in solution (Kd-part) was calculated by linear regression of the adsorbed amount against the free ion concentration, Cf, calculated as the difference between the measured equilibrium concentration in solution and the DOC-bound Cd concentration in solution, CDOC, estimated by WHAM.

A measure of the solubility of cadmium already present in the soil was obtained from the batch equilibrations performed without Cd addition:

[2]
where Kd-des is a desorption distribution coefficient (L kg–1) and Cini is the initial Cd content, here defined as the EDTA-extractable concentration (µg kg–1).

The significance of observed differences in Cd adsorption–desorption between SS and control treatments was determined in F tests, testing first for equality of variances, and then for equality of the slopes and intercepts derived from linear regression of log(Kd) against pH. Linear relationships between log(Kd) and pH can be expected from a theoretical consideration of surface complexation reactions (i.e., proton exchange on either hydroxyl groups on oxides and clay edges, or acidic groups on fulvic substances) and have also been demonstrated experimentally, both for the adsorption of added cadmium and the solubility of "native" cadmium (e.g., Gerritse and van Driel, 1984; Christensen, 1989; Lee et al., 1996; Janssen et al., 1997; McBride et al., 1997; Sauvé et al., 2000a).


   RESULTS AND DISCUSSION
TOP
 ABSTRACT
 INTRODUCTION
 MATERIALS AND METHODS
 RESULTS AND DISCUSSION
CONCLUSIONS
 REFERENCES
 
Soil Properties
Tables 1 and 2 show data on soil properties obtained from the first and second samplings. Within-treatment variation in soil properties of importance for cadmium sorption and solubility was rather small (Table 1, see also Bergkvist et al., 2003), with median coefficients of variation for pH, organic carbon content, and HNO3 and EDTA-extractable Cd of 1.3, 4.8, 9.3, and 11.4%, respectively. This is probably due to the limited areal extent of the field site, the small size of the plots (4 m2), and for the topsoil, the thorough homogenization resulting from hand-digging every year for the past 41 yr.


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  		Table 1. Basic soil properties (means, with coefficients of variation in parentheses).

 

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  		Table 2. Properties of the soil samples used in the batch experiments (second sampling).

 
The soil organic carbon content was approximately 70% larger (p < 0.0001) in the SS treatment topsoil compared to the control (Tables 1 and 2), due to the supply of organic carbon in sewage sludge and also to the higher biomass production that resulted from an improvement in soil fertility (Kirchmann et al., 1994; Bergkvist et al., 2003). The C to N ratio decreases with depth, indicating an increasing proportion of humified organic matter in the deeper soil layers. Despite reductions in soil organic carbon content with depth (e.g., from 1.5 to 0.58% in the control), the potential CEC varied very little, ranging between 212 and 246 mmolc kg–1. The HNO3 and EDTA-extractable Cd concentrations were approximately four times larger in the SS treatment topsoil compared to the control (p < 0.001 and p < 0.0001, respectively). The pH in the subsoil was approximately 6.4 in both treatments, while in the SS treatment topsoil, the pH was more than one unit lower compared to the control (p < 0.001) due to mineralization of organic compounds, accumulation of stable humus acids, and the higher biomass production (Kirchmann et al., 1996; Bergkvist et al., 2003).

Increased concentrations of noncrystalline and especially organically associated Fe and Al were found in the topsoil of the SS treatment (Table 3). Normalized to the soil organic carbon content, the pyrophosphate-extractable Fe concentration was seven times larger in the SS treatment topsoil (209 vs. 27 mg Fe g–1 organic C), and organically associated Al was five times larger (75 vs. 15 mg Al g–1 organic C). The concentrations of extracted Al and Fe fractions were similar in the upper subsoil layers of both treatments, indicating that sludge-borne Al and Fe compounds had not migrated from the topsoil into the subsoil. Topsoil Mn concentrations were not significantly different between treatments (Table 3).


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  		Table 3. Soil aluminium, iron, and manganese contents.{dagger}

 
Cadmium Speciation
Table 4 shows the input data for WHAM-W, derived from the measurements in equilibrium extracts without Cd addition. At similar pH, significantly larger (p < 0.001) DOC concentrations (22–35 mg DOC L–1) were found in the SS treatment topsoil compared to the control (7–14 mg DOC L–1). The DOC concentrations were also slightly, but significantly larger in the upper 7 cm of the subsoil in the SS treatment (p < 0.05). Table 4 also shows that the measurements on replicate samples in 2003 showed similar DOC and Cd concentrations, and that somewhat larger DOC concentrations were measured in the batch experiments in 2003 compared to 1999.


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  		Table 4. Measurements of cadmium and dissolved organic carbon (DOC) concentrations at controlled pH in zero addition equilibrations, and calculations of the proportion of Cd bound by DOC using the WHAM model.{dagger}

 
The calculations with WHAM-W for the control treatment suggested that less than 3% of Cd in solution was DOC complexed (Table 4). Even in the sludged soil with larger DOC concentrations, WHAM-W predicted that 83 to 98% of Cd in soil solution was present as Cd2+ (Table 4). Predictions of free ion concentrations using measured total solution concentrations and calculations with a speciation model like WHAM may be in error due to uncertainties and assumptions concerning the composition of Cd-complexing ligands and their associated stability constants (Sauvé et al., 2000b). Nevertheless, our calculations of the proportion of cadmium present as Cd2+ in solution are within the range of values measured on soil solutions in sludge-amended soils (Alloway and Tills, 1984; Knight et al., 1998; Holm et al., 1998).

Adsorption Distribution Coefficients
Figure 1 shows the relationships found between log(Kd-app) and pH. Cadmium sorption was strongly controlled by pH, with r2 values exceeding 92% for both treatments. The slopes of the regressions were not significantly different between treatments (p = 0.05) and are similar to those reported in the literature for Cd adsorption in a wide range of soils (Christensen, 1989; Lee et al., 1996; Gray et al., 1999) and are also within the range of H+–Cd2+ exchange ratios for humic acids (0.5–0.8) found by Kinniburgh et al. (1999). At any given soil pH, Cd adsorption was slightly but significantly weaker (i.e., smaller intercept at p < 0.05) in the SS treatment compared to the control (Fig. 1). For example, at pH 6, the Kd-app values were 194 and 149 L kg–1 in the control and SS treatment plots, respectively. The duplicate Kd-app values obtained in the second experiment in 2003 were similar to each other, and closely matched the log(Kd-app) vs. pH relationships obtained in 1999 (Fig. 1), demonstrating the excellent reproducibility of the experimental procedures and confirming that within-treatment variability in soil properties controlling cadmium sorption is small. In the second experiment, Kd-app in the SS treatment was significantly smaller (p < 0.05) than the control at pH 6.5, although the corresponding difference at pH 5.5 was not quite significant (p = 0.056).



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  		Fig. 1. Adsorption distribution coefficients (Kd-app) for cadmium as a function of pH. Open symbols are unreplicated measurements from 1999, while solid symbols represent measurements made on two replicates at two pH values in 2003.

 
Addition of sewage sludge to soil releases DOC into solution. Due to a high affinity for Cd, DOC may be one reason for the apparent reduction in the sorption affinity of the solid phase in the SS treatment (Neal and Sposito, 1986; Lamy et al., 1993; McBride et al., 1999). Utilizing WHAM-W calculations to account for the competition effects from DOC reduced the difference in Cd distribution coefficients (Kd-part) between the treatments (Fig. 2) , but the intercepts of the linear regression equations were still significantly smaller in the SS treatment (p < 0.01). However, the difference between treatments was relatively small (at pH 6, Kd-part is approximately 20% smaller in the SS treatment), and the predicted Kd-part values are sensitive to any uncertainties in the calculations of Cd complexation in WHAM-W (Kd-part values increase in proportion to the calculated DOC-complexed Cd fraction). Even allowing for Cd2+ fractions in solution at the lower end of the range of those measured in sludged soil (i.e., 60–80%; Alloway and Tills, 1984; Holm et al., 1998), our results suggest that the Cd sorption affinity of the immobile soil solid phase has at least not increased in the sludge-amended soil.



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  		Fig. 2. Adsorption distribution coefficients for cadmium on soil particles (Kd-part) as a function of pH.

 
Cadmium Solubility
Sludge amendment apparently had no effect on cadmium solubility, as no statistical difference (p = 0.05) was found between treatments in the slopes and intercepts of the log(Kd-des) vs. pH functions (Fig. 3) . The slopes of the functions (0.46 and 0.52) were not significantly different (p = 0.05) to those found in the adsorption experiments, and are also similar to values reported in the literature (e.g., Janssen et al., 1997). At any pH, topsoil Kd-des values were larger than the corresponding Kd-part values (e.g., at pH 6, Kd-des was 253 L kg–1 vs. estimated Kd-part values of 160 and 196 L kg–1 in the SS and control treatments, respectively), although it should be noted that the absolute values of Kd-des will depend on the definition of potentially available Cd (here EDTA-extractable). In contrast to the case with adsorption (Fig. 2), desorption coefficients in the upper 7 cm of the subsoils of both treatments were similar and were significantly larger than in the topsoils at any given pH (Fig. 3).



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  		Fig. 3. Desorption distribution coefficients (Kd-des) for cadmium as a function of pH.

 
Sludge "Protection"
In principle, some degree of sludge "protection" arises when the Kd value in sludge exceeds that of the native soil (McBride, 1995; Bergkvist and Jarvis, 2004). Complete protection, defined here as no increase (or even a decrease) in soil solution concentrations, occurs when the resulting increase in the Kd value of sludged soil is proportionally larger than the increase in bioavailable trace metal content in soil due to sludge loadings (Bergkvist and Jarvis, 2004). After 41 yr of biennial sludge application at Ultuna, the organic and inorganic sludge components now contribute significantly to the topsoil composition in the SS treatment (Tables 1 and 3). In total, 25 kg dry matter m–2 of sludge containing 40% inorganic components has been incorporated. The humic content has increased by approximately 70%, while the added inorganic precipitates (including 1.5 and 0.2 kg m–2 of Fe and Al oxides and hydroxides) now comprise approximately 3% of the total mass of the topsoil. In particular, the ammonium oxalate–extractable Fe concentration in the sludge-amended topsoil has more than doubled. However, no sludge "protection" occurred in our soil, as Cd sorption and solubility was unaltered, or even slightly reduced, in the sludge-amended soil compared to the control treatment (Fig. 1, 2, and 3). This implies that the Cd sorption affinity of the added biosolids (organic and inorganic fractions combined) must have been similar to the sorption affinity of the native soil. Our experiment was performed in an arable soil of high clay content, moderate organic matter content, and a pH near neutrality. These conditions result in a fairly high initial soil Cd sorption affinity (Kd = 105 L kg–1 for the control at pH 5.5). Hooda and Alloway (1994) also found no sludge "protection" effect in two soils with initial Kd values for Cd larger than approximately 500 L kg–1 at pH 6. Indeed, for one of these soils, Cd sorption 450 d after sludge application was weaker than in the control soil. Cline and O'Connor (1984) also showed that sludge application decreased Cd sorption for three alkaline soils with initial Kd values all larger than 450 L kg–1. In soils with much weaker initial Cd sorption, sludge additions might result in a permanently enhanced Cd sorption, especially if a significant proportion of this arises from the inorganic components in the sludge (Bergkvist and Jarvis, 2004). For example, Petruzzelli et al. (1997) found that Cd sorption in a sandy soil of very low CEC was still larger than a control soil 15 yr after sewage sludge applications ceased. Li et al. (2001) found that the adsorption Kd values of Cd on four long-term sludge-amended soils were between 28 and 210% times larger than the controls, with the organic and inorganic fractions contributing on average roughly equally to the enhanced sorption affinity. Hettiarachchi et al. (2003) demonstrated that the Kd value for Cd in a sandy loam soil amended with lime-treated digested and composted biosolids was two to four times larger than the control, due to increased organic carbon and Mn and Fe oxide contents. In these latter two studies, all five control soils had sorption Kd values for Cd less than approximately 60 L kg–1 at pH 5.5. Thus, variation in soil properties controlling initial Cd sorption may be one important reason for the apparently conflicting results of studies on long-term effects of sludge amendment on Cd sorption.

Apart from the Kd value of the original soil, the sorption affinity of sludge-derived organic and inorganic components, when mixed with the soil, represents the other side of the "sludge protection equation." The sorption Kd value for Cd on biosolids reported by Li et al. (2001) was approximately 1200 L kg–1 at pH 5.5, a value that was 20 times larger than the largest measured control soil Kd value. The strong "sludge protection" effect observed in the soils studied by Li et al. (2001) is therefore not surprising. Other authors also report sorption isotherms for Cd on sludge in this range (Riffaldi et al., 1983; Hooda and Alloway, 1994). However, in our study, the effective Kd value for Cd in the sludge applied during a 41-yr period to the SS treatment soil must have been considerably smaller, as it was apparently similar to the initial soil Kd value (approximately 105 L kg–1 at pH 5.5). The reason for this is not known, but it may be due to competition for sorption sites with other trace metals such as zinc (Christensen, 1987; Wilkins et al., 1998) or sludge-derived iron (Zachara et al., 1992; Pandeya and Singh, 2000). The latter hypothesis is supported by the fact that the pyrophosphate-extractable (i.e., organically associated) iron content was seven times larger in the SS treatment soil compared to the control (Table 3).


   CONCLUSIONS
TOP
 ABSTRACT
 INTRODUCTION
 MATERIALS AND METHODS
 RESULTS AND DISCUSSION
 CONCLUSIONS
REFERENCES
 
Cadmium solubility in a clay loam soil supplied biennially with sewage sludge during 41 yr was unaltered compared to the control, despite increases in organic matter and amorphous Fe oxide contents in the sludge-amended topsoil of 70 and 100%, respectively. Furthermore, Cd adsorption was slightly, but significantly (p < 0.01) weaker in the sludged soil, even after correcting for DOC complexation using the speciation model WHAM-W. The reasons for this are not clear, but it may relate to a relatively large initial sorption affinity for Cd in our soil, coupled with a reduction in the Cd sorption affinity per unit mass of organic carbon in the sludged soil, due to competition effects, primarily a blocking of the organic binding sites by Fe or trace metals such as zinc. In support of this hypothesis, the concentration of organically bound Fe was shown to be seven times larger in the sludge treatment compared to the control.

Our results contrast with those of two recent studies (i.e., Li et al., 2001; Hettiarachchi et al., 2003), although they are supported by others (O'Connor et al., 1983; Cline and O'Connor, 1984; Hooda and Alloway, 1994). Indeed, in principle, mixing sludge with soil may result in long-term increases or decreases in Cd sorption and solubility or no change at all, depending on the affinity for Cd of the sludge itself compared to the native soil and accounting for competition effects with other sludge-borne metals. This suggests that more research is needed to identify those situations where adequate sludge "protection" can be expected, and where it cannot, and in particular to clarify the role of competing trace metal cations and Fe oxides and hydroxides in controlling Cd sorption, solubility, and plant uptake in sludge-amended soils. The site-specific nature of the response means that it may be difficult to generalize about the long-term effects of sludge amendment. In a risk assessment context, models that simulate trace metal bioavailability in sludge-amended soils (e.g., Bergkvist and Jarvis, 2004) might then prove useful as diagnostic tools.


   ACKNOWLEDGMENTS
 
The Swedish Natural Science Research Council and SLU supported this research in the program "Biological Wastes in Circulation between Urban and Rural Areas." We wish to thank Jon-Petter Gustafsson (Royal Institute of Technology, KTH, Stockholm) for performing Fe, Al, and Mn analyses. We are also grateful to Ingmar Messing and Gudrun Sjöberg (Department of Soil Sciences, SLU) who assisted with statistical analysis and to Kenth Andersson, Lise Gustavsson, Gunilla Hallberg, and Gunilla Lundberg who assisted with soil analyses.


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


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