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Published in J. Environ. Qual. 33:902-910 (2004).
© ASA, CSSA, SSSA
677 S. Segoe Rd., Madison, WI 53711 USA

TECHNICAL REPORTS

Heavy Metals in the Environment

Assessment of the Use of Industrial By-Products to Remediate a Copper- and Arsenic-Contaminated Soil

Enzo Lombi*,a,b, Rebecca E. Hamonb, Gerlinde Wieshammera,c, Mike J. McLaughlinb and Steve P. McGratha

a Rothamsted Research, Agriculture and the Environment Division, Harpenden, Hertfordshire, AL5 2JQ, UK
b CSIRO Land and Water, PMB 2, Glen Osmond, SA 5064, Australia
c University of Natural Resources and Applied Life Sciences, Institute of Soil Science, Gregor Mendel Strasse 33, A-1180 Vienna, Austria

* Corresponding author (enzo.lombi{at}csiro.au).

Received for publication July 29, 2003.

   ABSTRACT
TOP
 ABSTRACT
INTRODUCTION
 MATERIALS AND METHODS
 RESULTS
 DISCUSSION
 CONCLUSIONS
 REFERENCES
 
Two water treatment sludges (WTS-A, WTS-B), two red muds (RM), and red gypsum (RG), all rich in iron oxy-hydroxides, were added to a soil highly polluted with As and Cu at 2% (w/w) to reduce metal bioavailability. Because the amendments increased soil pH to approximately 6, a lime treatment to the same pH and an unamended treatment were included for comparison. All the amendments had significant positive effects on the soil microbial biomass and growth of ryegrass (Lolium multiflorum Lam. cv. Avance), but only WTS-A improved lettuce (Lactuca sativa L. cv. Tom Thumb) growth. The mineralization of added ammonium nitrogen was not significantly affected by the treatments, while a physiologically based extraction test (PBET) showed that bioaccessibility of As was low (<5%) and decreased only in the WTS-A treatment. Concentrations of As in soil pore water and extractable As only decreased in the WTS and RG treatments. In contrast, Cu concentrations in soil pore water and extractable Cu decreased in all treatments, by more than 84% in the WTS, RM, and RG treatments. Non-isotopically exchangeable As and Cu were present in colloids in the soil pore water. Untreated soil had <4% isotopically exchangeable As and this decreased by approximately 50%, with WTS, RM, and RG. The labile Cu pool represented a large proportion (34%) of the total Cu pool, and the isotopically exchangeable and soluble Cu were strongly correlated with soil pH. Acidification of the treated soils showed that the labile As and Cu both increased in the treated soils compared with untreated soils. The significance of the treatment effects on soil fertility and potential off-site transport of As and Cu to ground water are discussed.

Abbreviations: E, labile pool of arsenic and copper obtained from the filtrates • Er, labile pool of arsenic and copper obtained from the resin-purification step • PBET, physiologically based extraction test • RG, red gypsum • RM, red mud • WTS, water treatment sludge


   INTRODUCTION
TOP
 ABSTRACT
 INTRODUCTION
MATERIALS AND METHODS
 RESULTS
 DISCUSSION
 CONCLUSIONS
 REFERENCES
 
CONCENTRATIONS OF HEAVY metals and metalloids in soils exceeding permissible limits according to current standards have been identified in many regions of the world. Arsenic and Cu are widespread contaminants as a result of both human activities and geogenic factors. For instance, As is the second most common inorganic contaminant after Pb at Superfund sites in the USA, being present at 41% of the sites (USEPA, 1997). Perhaps the most important of the anthropogenic sources of As is that associated with mining and smelting activities. Arsenic is contained in Cu, Pb, Co, and Au ores in amounts ranging from 20 to 30 g kg–1 in Cu ores and up to 110 g kg–1 in some Au ores (Azcue and Nriagu, 1994). Smelting of these ores can lead to complicated situations where mixed pollution with anionic As and cationic metals, such as Cu, is present. A similar scenario is present when wood preservatives containing chromated copper arsenates (CCA) are used.

While various remediation technologies are available for cationic metals, feasible treatment options for As-polluted environments are limited (Lombi et al., 2000). Because As in water and soil systems is present in anionic forms (Sadiq, 1997), conventional remediation technologies for heavy metals often produce undesired effects when employed at As-polluted sites. For example, phosphates, organic matter, and lime have been used to decrease the mobility and toxicity of heavy metals (Cunningham and Berti, 2000). However, these amendments may not be effective in the case of As-contaminated soils and may even result in increased mobility and bioavailability of this metalloid because of competition between the As oxyanion with hydroxyl ions, phosphate, and organic ligands for adsorption sites (Sadiq, 1997; Boisson et al., 1999a; Grafe et al., 2001, 2002).

Oxy-hydroxides of Fe and Al have been identified as primary sinks for As in soils (Lombi et al., 1999). Therefore, adsorption and immobilization of As in soil may be achieved by furnishing additional quantities of these As sinks to the soil. Boisson et al. (1999b) reported that application of steel shot to a contaminated soil, at a dose of 10 g kg–1 of soil, was effective in decreasing the mobility of As. Similarly, application of steel shot to an As-enriched garden soil was shown to decrease As contents in vegetables (Vangronsveld, 1998). Iron oxy-hydroxides have also been synthesized in situ by applying ferrous sulfate to a contaminated soil (Artiola et al., 1990; Voigt et al., 1996). Since Fe oxides are also able to bind heavy metals, the use of Fe-rich materials could be appropriate to treat soils contaminated with both As and cationic metals.

Our understanding of remediation technologies to simultaneously treat anionic and cationic inorganic contaminants is far from complete. Here we assess the risk reduction caused by the use of various Fe-rich industrial by-products to remediate a soil contaminated with both As and Cu. This risk assessment focuses on environmentally relevant endpoints (microorganisms, plants, and humans) and pathways of exposure (ground and pore water contamination, direct ingestion). An important microbial soil function (N mineralization) and soil microbial biomass were chosen as microbial endpoints. Two plant species relevant in terms of human and animal nutrition were considered. The human risk related to direct ingestion of soil was estimated using a physiologically based extraction test (PBET; Ruby et al., 1999). The potential for contamination of ground water and colloid-facilitated transport of contaminants were investigated in soil pore waters. To gain information on the mechanisms of As and Cu immobilization, a recently developed isotopic dilution technique was employed. This method was also used to assess the potential risk related to acidification of the remediated soils.


   MATERIALS AND METHODS
TOP
 ABSTRACT
 INTRODUCTION
 MATERIALS AND METHODS
RESULTS
 DISCUSSION
 CONCLUSIONS
 REFERENCES
 
Soil and Amendments
The soil used in the trial was collected from an agricultural grassland located 200 m from the Devon Great Consols Mine, Devon, UK (50°32'15'' N, 4°13'23'' W). This mine covers 68 ha of land in the southern part of the Tamar Valley. This region was the world's largest producer of Cu and As during the 19th century. The principal As-containing ore is arsenopyrite and Cu is largely present as chalcopyrite (Goodridge, 1964). The soil was collected from the surface (0–20 cm) of the grassland and the source of contamination is likely to be both geogenic and anthropogenic due to the smelting of the ores at the mining site. The soil samples were homogenized and characterized for pH (1:2.5 suspension of soil and water), and total C and N were determined using a LECO (St. Joseph, MI) CNS-2000 analyzer. Total concentrations of heavy metals in the soils were determined using inductively coupled plasma–atomic emission spectroscopy (ICP–AES) (Fisons Accuris; Thermo Electron, Waltham, MA), following microwave digestion using HNO3 and H2O2. Arsenic in the digest was determined using atomic absorption spectrometry coupled with a hydride generation system (FIAS 400; PerkinElmer, Wellesley, MA). Total concentrations of As and metals were also analyzed in certified soil standards (Montana soil, NIST 2711; National Institute of Standards and Technology, Gaithersburg, MD) to ensure quality control. The sand, silt, and clay contents were determined using the pipette method (Avery and Bascomb, 1982). Subsamples (3 kg) of the soil were mixed with one of five Fe-rich industrial by-products. These soil amendments were added at 2% (w/w). The amendments included two water treatment sludges collected from water treatment plants in Germany (WTS-A and WTS-B), two red muds (from Hungary [RM-H] and from the United Kingdom [RM-UK]), and a red gypsum (RG) collected from a Ti-oxide factory in the UK. In the case of RM, an addition rate of 2% was reported to decrease metal mobility and toxicity in a previous study (Lombi et al., 2002a, 2002b). All the soil amendments were characterized as described above for the soil. Since RM is an alkaline material and the pyritic soil used had a low pH, Ca(OH)2 was added to the remaining treatments (except the untreated control) to obtain a pH similar to that of the RM-treated soil (pH = approximately 6). Soil mixed with the different amendments was used to fill 1-kg pots. Also included in the experimental design were untreated soil and a pH control (lime treatment) where Ca(OH)2 was used to raise the soil pH to 6.0. Three replicates per treatment were prepared.

Plant Growth Assessment
The pots were placed in a glasshouse, watered, and left to equilibrate for 10 wk. After this period two crops of ryegrass were grown in the pots, followed by lettuce. The durations of the cropping cycles were 5, 3, and 6 wk for the first and second cut of ryegrass and for lettuce, respectively. Ryegrass was chosen because of its use as animal fodder. Lettuce is a sensitive crop relevant for human exposure. Both crops were directly sown in the pots. Each crop was fertilized with a nutrient solution as reported in Lombi et al. (2002a). Between the ryegrass and lettuce crop, the soil in each pot was mixed and incubated for 2 wk at 60% water-holding capacity. Plants were harvested by cutting at the soil surface, washed thoroughly with deionized water, and oven-dried (at 60°C for 48 h), and the shoot dry weight was recorded. Shoots of both crops were finely ground and digested (0.5 g) using a mixture of HNO3 and HClO4 (Zhao et al., 1994). The concentrations of heavy metals and As in the shoots were determined as described above. Similarly to soil, total concentrations of As and metals were also analyzed in certified plant standards (spinach leaves, NIST 1570a) as a quality control procedure.

Microbial Endpoints
An important microbially driven soil function, N mineralization, and a measure of the total soil microbial biomass were used as microbial endpoints. Immediately following the final harvest, soils were subsampled for determination of substrate-induced N mineralization and soil microbial biomass. The substrate-induced N mineralization test was performed over a period of 4 wk following the OECD Guideline 216 (Organisation for Economic Co-Operation and Development, 2000). The substrate added was finely ground bean plants. Soil microbial biomass was measured following the fumigation–extraction method (Wu et al., 1990).

Arsenic Bioaccessibility
Soil subsamples for this test were collected at the end of the pot trial. Potential bioaccessibility of As to humans was assessed using the PBET developed by Ruby et al. (1999). This method was designed to mimic solubility-limiting conditions in a child's digestive tract.

Assessment of Potential for Water Contamination
Potential contamination of surface or ground water was assessed by measuring soil pore water concentrations of As and Cu and by chemical extractions. These measurements are relevant both in terms of exposure of soil organisms through pore water and in view of potential leaching of As and Cu to ground water. Colloidal-facilitated transport of contaminants to ground water was also investigated using an isotopic dilution technique described below. Soil pore water (10 mL) was collected from each pot after the final harvest, using nylon-coated soil moisture samplers (Rhizosphere Research Products, Wageningen, the Netherlands). Metals and As concentrations were measured as described above. Subsamples of soil, also collected after the final harvest, were air-dried and ground. Extractability of Cu was determined in 1 M NH4NO3 following DIN Standard 19730 (Deutsch Institut fur Normung, 1995). Arsenic extractability was assessed by 0.05 M (NH4)2SO4 extraction as described by Wenzel et al. (2002). These two extractants were chosen because extractant thresholds to meet water quality criteria are available for Cu and As (Pruess, 1994; Wenzel et al., 2002).

Potential Risk Due to Re-Acidification of Remediated Soils and Mechanisms of Immobilization
These aspects were investigated using isotopic dilution techniques. Their use enables labile As and Cu (soluble plus reversibly sorbed) to be distinguished from forms that are nonreactive (non-isotopically exchangeable). Investigation of the stability to acidification of the immobilized Cu and As in WTS-A- and lime-treated soils was performed coupling a conventional isotopic dilution technique with a stepwise acidification procedure as described in Hamon et al. (2002) and Lombi et al. (2003). The WTS-A treatment was selected because it gave the largest decrease overall in hazard as assessed with the biological indicators described above. Lime was chosen to investigate whether the effect of WTS-A was solely related to changes in pH. We also included a resin-purification step (as described in Hamon and McLaughlin, 2002 and Lombi et al., 2003) to assess the presence and exchangeability of As and Cu associated with colloids (<0.2 µm) in the solution phase. This information is essential for evaluating the potential for colloidal-facilitated transport of Cu and As.

Subsamples (2 g) of the treated soils were placed in centrifuge tubes with deionized water and two drops of toluene. In the case of the WTS-A and lime treatments appropriate aliquots of dilute HCl were also added to obtain a series of five pH levels with the lowest pH being <4.0. The final soil to solution ratio was 1:10. The soil suspensions were equilibrated for 4 d in an end-over-end shaker. The pH was then measured (PHM93; Radiometer, Brønshøj, Denmark) and the samples were spiked with 50 µL of solution containing 73As (600 kBq mL–1) or 64Cu (60 MBq mL–1). In the case of 64Cu, the activity added to the samples was large due to the short half life (12 h) and low efficiency of {gamma} counting for this isotope. Samples were then returned to the shaker and left to equilibrate for 2 d in the case of the 73As-spiked soils and 1 d for samples containing 64Cu. The samples were then centrifuged and filtered through 0.2-µm cellulose acetate filters (Sartorius, Goettingen, Germany). Ten milliliters of the filtrates were placed in tubes containing 100 mg of Chelex 100 resin (Bio-Rad Laboratories, Hercules, CA) or a 5-cm2 anion resin strip (BDH, Chicoutimi, QC, Canada) for the samples spiked with 64Cu and 73As, respectively (Lombi et al., 2003; Hamon and McLaughlin, 2002). The samples were equilibrated for 2 h with end-over-end shaking. The solution was then removed and the resin rinsed twice with distilled water. Copper was eluted from the Chelex 100 resin using 5 mL of 0.5 M HNO3. Arsenic was eluted from the anion strip using 5 mL of 0.5 M NaCl acidified to pH 1 with HCl. Total concentrations of As and Cu in the remaining aliquots of the filtrates and in the eluants were measured as described above. Activities of 73As and 64Cu were determined using {gamma} spectrometry (1480 Wizard; Wallac, Turku, Finland). All analyses were performed in triplicate and included blanks as well as solutions spiked with radioisotope so that the total amount of radioisotope added could be determined. The labile pool of As and Cu obtained from the filtrates (E) and from the resin-purification step (Er) were determined as reported in Hamon et al. (2002) using the formula:

[1]
where Csol (mg mL–1) is the concentration of non-radioisotopic metal in the filtrates (E values) or elutes (Er values), C*sol is the activity of radioisotope in the filtrates or elutes (Bq mL–1), R is the total amount of radioisotope that was added to each sample (Bq mL–1), and V/W is the ratio of solution to sample, which in this case was 10 mL g–1.

Comparison of the E values with the Er values allows quantification of the fraction of As and Cu associated with colloids that is not isotopically exchangeable. At the same time, the resin-purification step overcomes the possible interference of colloids in the determination of E values (Lombi et al., 2003).

Statistical Analyses
Data were analyzed using analysis of variance (ANOVA) or correlation analyses. Least significant difference (LSD) was used for comparison between the treatment means. All statistical analyses were performed using the Genstat 5 package (Genstat 5 Committee, 1993).


   RESULTS
TOP
 ABSTRACT
 INTRODUCTION
 MATERIALS AND METHODS
 RESULTS
DISCUSSION
 CONCLUSIONS
 REFERENCES
 
Water treatment sludge is a by-product of the water industry and is rich in Fe due to precipitation of oxy-hydroxides during purification of surface or ground water. Red mud and red gypsum are by-products generated during the production of alumina and Ti oxides, respectively. All the soil amendments were rich in Fe oxy-hydroxides with a total content ranging from 17 to 33% (w/w) (Table 1). Red muds were alkaline whereas the other soil amendments had a neutral pH. Concentrations of As, Cu, and Cd in the soil were much larger than in any of the amendments. The soil pH increased by approximately 1 to 1.5 units in all the amended treatments in comparison with the control (Table 2).


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  		Table 1. Principal chemical and physical characteristics of the soil and soil amendments used.

 

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  		Table 2. Soil pH, extractable and soil pore water Cu and As, physiologically based extraction test (PBET) As, and soil microbial biomass determined at the end of the pot trial.{dagger}

 
Plant Growth Assessment
With the exception of RM, all the amendments significantly (P < 0.001) increased the shoot biomass of ryegrass at both harvests in comparison with the untreated control (Fig. 1a) . In particular, plants grown in the WTS-A-treated soil had the largest shoot biomass at the second harvest (more than nine times the biomass of the control). Lettuce was found to be more sensitive to metal toxicity than ryegrass and grew poorly in all the soils. Only lettuce grown in the WTS-A had a significantly (P < 0.001) larger shoot biomass than plants grown in any other treatment or in the control soil. Arsenic concentrations in ryegrass were greater than in lettuce (Fig. 1b). Arsenic concentrations were greater than the normal range observed in plants (1–1.7 mg kg–1; Kabata-Pendias and Pendias, 1992) but within the concentration range observed in plants growing at As-contaminated sites (1–35 mg kg–1; Brandstetter et al., 2000). Even though some of the treatments seemed to decrease As accumulation in ryegrass in comparison with the control, these differences were not statistically significant. Arsenic concentrations in lettuce grown in the WTS, RM-H, and RG treatments were significantly (P < 0.05) smaller than in the other treatments. With the exception of the first ryegrass harvest in the RG treatment, Cu concentrations in plants in all the treatments were significantly (P < 0.001) smaller than in the control (Fig. 1c). Combining the data regarding plant biomass and metal concentration it is also possible to obtain indications regarding the potential Cu and As transfer in the food chain through herbivory.



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  		Fig. 1. Shoot dry biomass of lettuce and ryegrass, and As and Cu concentrations in dry shoots of plants grown in the untreated control soil and in the amended soils. Error bars represent ±SE of three replicates. RG, red gypsum; RM-H and RM-UK, red mud from Hungary and the United Kingdom, respectively; WTS, water treatment sludge.

 
Microbial Endpoints
Nitrogen mineralization was assessed over a period of 4 wk (Fig. 2) . At the end of this period no statistically significant differences were observed between the treatments. These results are surprising because the nitrification rate typically increases with increasing soil pH. Soil microbial biomass, measured at the end of the pot trial, was significantly (P < 0.01) increased in the lime- and WTS-A-amended soils in comparison with all the other treatments (Table 2).



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  		Fig. 2. Nitrogen mineralization measured after 7, 14, and 28 d in the untreated control soil and in the amended soils. Error bars represent ±SE of three replicates. RG, red gypsum; RM-H and RM-UK, red mud from Hungary and the United Kingdom, respectively; WTS, water treatment sludge.

 
Arsenic Bioaccessibility
The PBET was used to replicate the solubility-limiting conditions in a child's digestive tract (Ruby et al., 1999). The only significant (P < 0.001) decrease in PBET As was observed in the WTS-A treatment (Table 2). In all the other treatments PBET As did not change or increased slightly (lime treatment).

Assessment of Potential for Water Contamination
In soil, exposure of plants and microorganisms to metals occurs primarily through the aqueous phase (Plette et al., 1999). Furthermore, contaminants in the soil pore water may be leached to ground water. To assess the effect of the remediation treatments in terms of potential for water contamination, Cu and As concentrations in soil pore water were measured at the end of the pot trial. Copper concentrations were significantly (P < 0.001) decreased in all the amended soils in comparison with the untreated control (Table 2). The As concentration in soil pore water was extremely large in the untreated control (1.8 mg L–1) and was not significantly decreased in the limed and RM-treated soils. All the other treatments significantly (P < 0.01) decreased pore water As concentrations. Among these treatments, WTS-A produced the largest decrease in pore water As (a decrease of 75% in comparison with the control; Table 2).

Ammonium nitrate extraction (DIN Standard 19730; Deutsch Institut fur Normung, 1995) is used in Germany to assess action and threshold values for metals in soils to protect water quality. Extractability of Cu was lower than the threshold of 1000 mg kg–1 proposed by Pruess (1994) for protection of water quality. All soil amendments significantly decreased Cu extractability in comparison with the control (P < 0.001; Table 1). In particular, WTS-A, RM-H, and RG were the three most effective treatments with a decrease of up to 84% compared with the untreated soil. Ammonium sulfate extraction has been used to predict leachability of As under field conditions (Wenzel et al., 2002). With the exception of lime, all amendments significantly (P < 0.001) decreased the extractability of As in comparison with the untreated soil; with a 55% decrease observed in the WTS-A treatment. However, extractable As was much larger than the value of 0.62 mg kg–1 proposed by Wenzel et al. (2002) for freshwater protection.

A resin method, in combination with an isotopic dilution technique, was used to assess the presence of nonlabile Cu and As associated with soil colloids in the solution phase. This assessment is relevant in terms of ground water protection because colloids may facilitate the transport of strongly bound contaminants to ground water. As discussed in Lombi et al. (2003) and Hamon et al. (2002), the resin used in the purification step provides a sink for exchangeable As and Cu; in contrast, non-exchangeable metals associated with colloids in the solution phase are not adsorbed by the resin. A comparison between the labile pool measured with (Er value) or without (E values) the resin-purification step allowed us to assess whether non-isotopically exchangeable metals were present on <0.2-µm colloids in the solution phase. The Er values were generally less than or equal to E values (i.e., E/Er > 1), indicating that non-isotopically exchangeable Cu and As were present on colloids in the solution phase. The E values tended to be larger than Er values with increasing pH for both As and Cu (Fig. 3a and 3b , respectively).



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  		Fig. 3. Relationship between the ratio of labile pool of arsenic and copper obtained from the filtrates (E) and that obtained from the resin-purification step (Er), and pH of the soil suspensions. All replicate and treatment data are reported.

 
Potential Risk Due to Re-Acidification of Remediated Soils and Mechanisms of Immobilization
Arsenic lability (Er value) was calculated using Eq. [1] assuming that all As in the soil was present as As(V). This assumption was verified by speciation of the untreated soil extracts by high performance liquid chromatography (HPLC)–inductively coupled plasma (ICP)–mass spectrometry (MS). The Er values, which are free from the interference of colloids, are used to assess changes in lability due to remediation treatments or to changing pH conditions. The results showed that only a small proportion of the total As was labile in the soil (Fig. 4a) . For instance, in the untreated soil only 3.7% of the total soil As was isotopically exchangeable; WTS-B, RM, and RG further decreased As lability in comparison with the control. The acidification tests conducted with the lime and WTS-A treatments indicated that in both cases acidification increased As lability. Interestingly, these two treatments caused the Er values to be larger than in the control treatment at the same pH. The soluble As, determined as the As sorbed by the resin from the water extract during Er calculations, is shown in Fig. 4b. The untreated soil and soils treated with RG or WTS had <5 mg kg–1 soluble As, even when the WTS treatment was acidified. In contrast, As solubility in the soil treated with alkaline materials (lime and RM) was significantly greater.



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  		Fig. 4. (a) Labile As pools and (b) As sorbed by the resin from the water extract (Er) as a function of soil pH and treatments. Arrows indicate the control treatment. Error bars represent ±SE of three replicates. RG, red gypsum; RM-H and RM-UK, red mud from Hungary and the United Kingdom, respectively; WTS, water treatment sludge.

 
Isotopically exchangeable Cu generally represented a much larger fraction of the total soil Cu than found for As. For instance, in the control soil 34% of the total soil Cu was labile (Fig. 5a) . Copper lability showed a linear increase, independent of treatment, as soil pH decreased. Similarly, a strong relationship was observed in all treatments between soil pH and Cu extracted with the resin during Er determination (Fig. 5b).



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  		Fig. 5. (a) Labile Cu pools and (b) Cu sorbed by the resin from the water extract (Er) as a function of soil pH and treatments. Arrows indicate the control treatment. Error bars represent ±SE of three replicates. RG, red gypsum; RM-H and RM-UK, red mud from Hungary and the United Kingdom, respectively; WTS, water treatment sludge.

 

   DISCUSSION
TOP
 ABSTRACT
 INTRODUCTION
 MATERIALS AND METHODS
 RESULTS
 DISCUSSION
CONCLUSIONS
 REFERENCES
 
The positive effect of soil amendments on plant growth may be attributed to an increase in pH of the amended soils and to a decrease in Cu and As phytoavailability as estimated with pore water analyses and soil extractions (Table 2). Copper concentrations in plants grown in the control treatment (up to 30 mg kg–1) could not be considered as severely phytotoxic. However, shoot Cu and As concentrations are not good indicators of phytotoxicity because root to shoot translocation of these elements is often highly regulated and Cu toxicity first acts at root level. Furthermore, the differential response observed in the case of lettuce, the most sensitive of the two species grown, between different amendments suggests that the mechanism of action of WTS-A was not simply related to a pH effect or decreases in available Cu and As (Fig. 1). In fact, other soil amendments increased soil pH and decreased Cu extractability to a similar degree as WTS-A but their effect in terms of plant growth was smaller (Table 2, Fig. 1). Arsenic concentrations in ryegrass and lettuce both exceeded recommended concentrations in fodder crops and foodstuffs. For instance, in Austria the maximum limit for As in fodder crops is set at 2.0 mg kg–1 (Bundesgesetzblatt fuer die Republik Oesterreich, 1994), while in Australia the maximum permissible limit for cereals is 1.0 mg As kg–1 on a fresh weight basis (National Food Authority, 1993). Another point that is worth mentioning is that P concentrations in the plants were low (data not shown). Maintaining an adequate plant P nutrition during remediation of As-contaminated soils is problematic due to the similarities between arsenate and phosphate. Iron-rich materials that are efficient in binding As will also decrease P availability. High levels of labile Cu in the soil can also decrease available P due to the formation of insoluble Cu phosphate. On the other hand, an increased P fertilization to supply this nutrient to the plants could mobilize As from the soil and pose a risk for ground water contamination.

Soil microbial biomass was increased only in the lime and WTS-A treatments. This increase may be due to a direct decrease of metal toxicity to microorganisms. In the case of the WTS-A-treated soil, microbial growth may also have been stimulated by an increase in plant biomass, and consequently an increase in root exudation of C substrates. Confounding factors such as differences in soil pH, acclimation of soil microbial processes, and tolerance of microbial populations exposed for long periods in field-contaminated soil may explain why N mineralization appeared to be unaffected by soil treatments. It has been demonstrated that microbial populations of metal-contaminated sites tolerate larger metal concentrations than those of uncontaminated soils (Diaz-Ravina and Baath, 2001; Baath et al., 1998; J. Rusk, unpublished data, 2003).

Oral bioaccessibility, estimated using the PBET method, was small (5% of the total As) in the untreated soil. This is comparable with a prediction of 15.3% made using a steady state model of As(V) bioaccessibility developed by Yang et al. (2002). This model is based on Fe content of the soil and pH. The PBET As was significantly decreased only in the WTS-A treatment. It is probable that the Fe oxy-hydroxides contained in this amendment were extremely effective in adsorbing and retaining As. Yang et al. (2002) reported that soil pH and Fe oxide content are the most important soil properties controlling bioaccessibility of As.

The soil amendments investigated generally decreased the potential for water contamination by Cu and As. The ability of Fe-rich material to sorb heavy metals and As is well documented (e.g., Lindsay, 1979; Bruemmer et al., 1988; Lombi et al., 1999; Wenzel et al., 2001). Boisson et al. (1999b) reported that steel shot efficiently decreased the extractability of heavy metals, including Cu and As, from contaminated soils. Mench et al. (2003) studied the effect of various soil amendments in terms of As and metal mobility when applied to a barren Au mine spoil in Portugal. Steel shot, in combination with compost, produced the best results in terms of decreasing As and metal leaching. Copper concentrations in soil pore water and extractable Cu were decreased by all soil amendments (Table 2). Copper concentrations in NH4NO3 extracts were much lower than the proposed threshold values for protection of water quality (1000 mg kg–1; Pruess, 1994). Soil pore water As and extractable As were only decreased by Fe-rich amendments. This can be explained by the fact that the increased mobility of the anionic As ion with pH (Sadiq, 1997) is counterbalanced by the addition of sorption sites on the Fe oxy-hydroxides added with the Fe-rich materials. The WTS-A treatment produced the largest decrease in extractable As and soil pore water As. However, As concentrations in soil pore water were still far greater than the 10 µg L–1 value recommended by the World Health Organization for drinking water and adopted by a large number of countries (World Health Organization, 1994). The differences in terms of metal and As extractability and soil pore water concentrations between Fe-rich materials were probably due to differences in the mineralogical composition, surface area, organic matter content, and ability to mobilize soil organic matter from these materials. For instance, Lombi et al. (2002a) reported that RM, due its high alkalinity and NaOH content, increased dissolved organic carbon (DOC) in soil pore waters. Grafe et al. (2001)(2002) demonstrated that DOC can decrease adsorption of As(III) and As(V) on goethite.

Comparison of the E value data obtained with the resin-purification step (Er values) and with E values determined conventionally (E values) indicated that in the case of both As and Cu, non-isotopically exchangeable forms of trace elements are associated with <0.2-µm colloids in the solution phase. The difference between E and Er values increased with pH. This agrees with previous observations by Lombi et al. (2003). The fact that non-isotopically exchangeable Cu and As may be present in the mobile colloidal fraction of soils is of importance in terms of bioavailability and potential for off-site contaminant transport. Nonlabile metals are considered to be of low chemical and biological significance because they are not readily bioavailable (Smolders et al., 1999). In a column study, Grolimund et al. (1996) showed that colloidal-facilitated transport of strongly sorbing contaminants may be significant in natural porous media such as soil. We have demonstrated that assuming that all the metals in, or on, the mobile colloids are readily bioavailable is erroneous. Nonlability of colloidal metals may be important in several ways for contaminant transport and ecosystem effects, and as yet these issues remain unexplored. If significant amounts of colloidal metals are not in equilibrium with the soil solution, then transport of metals through reactive media such as soils may be underestimated. Similarly, risks of colloidal metals to aquatic biota may be overestimated since a significant amount of the measured metal in solution is essentially non-bioavailable, provided no change in lability occurs over time or space. The key factors controlling the transformation of nonlabile colloidal metal into labile forms remain unstudied.

Similarly to results for PBET, only 3.7% of the total As was isotopically exchangeable (labile). This can be explained by the presence of As in sparingly soluble forms, such as arsenopyrite, in the soil (Goodridge, 1964). A similar conclusion was drawn by Tye et al. (2002) who reported that labile As was between 1.4 and 19% of the total As in soils from two mining sites in Malaysia and the UK. Soil amendments produced small decreases in As lability (Fig. 4a). When the WTS-A-treated and limed soils were acidified down to pH 4.5, As lability increased slightly. After this point As lability seemed to stabilize (unfortunately only one data point is available at pH of <4.5). Tye et al. (2002) reported that As lability in soils collected at a mining site tended to increase with decreasing pH. However, it is interesting to note that at parity of pH with the untreated soil, As is more labile in the re-acidified soil samples treated with lime and WTS-A than in the control soil. A similar result was reported for Zn in a re-acidified lime-treated soil by Hamon et al. (2002). We hypothesize that when lime or WTS are added to the soil an increasing amount of As is sorbed by carbonates and Fe oxides, which depletes As in the solution phase. As a consequence non-exchangeable As is released by the soil solid phase and accumulates in the added carbonates and Fe oxides in a pH-reversible manner. When the soil is re-acidified this As becomes isotopically exchangeable and increases As lability to values larger than in the unamended soil. This hypothesis is supported by the results for As solubility reported in Fig. 4b. In the limed soil, acidification increased As solubility most likely due to desorption of As associated with carbonates. In the case of the WTS-treated soil, As solubility remained constant probably because the increased lability of As (Fig. 4a) was counterbalanced by the increased number of exchange sites on the Fe and Al oxides present in the soil and added with the sludge. Also, the solubility of As in soil amended with alkaline materials (lime and RM) was greater than in soil treated with Fe-rich but neutral materials such as RG and WTS.

Unlike As, the labile Cu pool represented a large proportion (34%) of the total Cu pool. In the case of Cu, the Er values and Cu solubility were strongly correlated with soil pH, independent of the treatments (R2 = 0.84 and 0.80, respectively). This is in agreement with the findings of Lombi et al. (2003), who reported a similar pattern in Cu Er values when contaminated soils amended with lime, beringite, and RM were re-acidified. Also, re-acidification of the treated soils produced a much more marked increase in Cu Er value than that observed for As.


   CONCLUSIONS
TOP
 ABSTRACT
 INTRODUCTION
 MATERIALS AND METHODS
 RESULTS
 DISCUSSION
 CONCLUSIONS
REFERENCES
 
The evaluation of the remediation efficiency of different materials should be based on a comparative risk assessment of the treatments taking into account relevant pathways of exposure, different environmental endpoints, and the different sensitivity of such endpoints. For instance, difference in sensitivity and response to treatments were evident both between different classes of organisms (plants, microorganisms) and within the same class depending on the endpoint considered (N mineralization vs. soil microbial biomass). The results suggested that some Fe-rich industrial by-products could be used for remediation of soils co-contaminated with heavy metals and anionic metalloids. However, significant differences exist between materials even within the same class of by-product due to differences in the treatment or process used at the industrial level and the different nature of the minerals and waters used in these processes. These differences most likely result in large variability in terms of composition and mineralogy of the by-products. Here, WTS was the most effective amendment in terms of enhancing plant and microbial growth, decreasing metal and As mobility, and diminishing bioaccessible As. However, isotopic dilution techniques in combination with a step-wise acidification procedure showed that the decrease in metal lability observed with this treatment is not stable if re-acidification was to occur. While of significance for Cu, this would be less of a problem for As due to an increase in As sorption on soil oxides at low pH. The isotopic dilution technique coupled with a resin-purification step confirmed previous reports (Lombi et al., 2003) that non-isotopically exchangeable metals and metalloids may be associated with mobile colloids in soil. This must be considered in terms of the potential to enhance off-site contaminant transport, especially in this case As, to ground water.


   ACKNOWLEDGMENTS
 
This work was funded through a grant from the International Lead Zinc Research Organisation. Rothamsted Research receives grant-aided support from the UK Biotechnology and Biological Science Research Council. The authors would also like to thank the Australian Institute for Nuclear Science and Engineering for providing financial assistance (Grant no. 02/087) to enable purchase of the 64Cu radioisotope.


   REFERENCES
TOP
 ABSTRACT
 INTRODUCTION
 MATERIALS AND METHODS
 RESULTS
 DISCUSSION
 CONCLUSIONS
 REFERENCES
 


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JEQ 2004 33: 799-804. [Full Text]  




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