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Journal of Environmental Quality 30:1222-1230 (2001)
© 2001 American Society of Agronomy, Crop Science Society of America, and Soil Science Society of America

TECHNICAL REPORT
Heavy Metals in the Environment

Chemical Immobilization of Lead, Zinc, and Cadmium in Smelter-Contaminated Soils Using Biosolids and Rock Phosphate

N.T. Basta*,a, R. Gradwohlb, K.L. Snethen and J.L. Schroder

a Dep. of Plant and Soil Sciences, Oklahoma State Univ., Stillwater, OK 74078
b CRC Environmental, Inc., Tulsa, OK

* Corresponding author (bastan{at}okstate.edu)

Received for publication May 9, 2000.

   ABSTRACT
TOP
 NOTES
 ABSTRACT
INTRODUCTION
 METHODS AND MATERIALS
 RESULTS AND DISCUSSION
 SUMMARY AND CONCLUSIONS
 REFERENCES
 
Chemical immobilization, an in situ remediation method where inexpensive chemicals are used to reduce contaminant solubility in contaminated soil, has gained attention. We investigated the effectiveness of lime-stabilized biosolid (LSB), N-Viro Soil (NV), rock phosphate (RP), and anaerobic biosolid (AB) to reduce extractability and plant and gastrointestinal (GI) bioavailability in three Cd-, Pb-, and Zn-contaminated soils from smelter sites. Treated (100 g kg-1 soil) and control soils were incubated at 27°C and -0.033 MPa (-0.33 bar) water content for 90 d. The effect of soil treatment on metal extractability was evaluated by sequential extraction, on phytoavailability by a lettuce bioassay (Lactuca sativa L.), on human GI availability of Pb from soil ingestion by the Physiologically Based Extraction Test. The largest reductions in metal extractability and phytoavailability were from alkaline organic treatments (LSB and NV). Phytotoxic Zn [1188 mg Zn kg-1 extracted with 0.5 M Ca(NO3)2] in Blackwell soil (disturbed soil) was reduced by LSB, NV, and RP to 166, 25, and 784 mg Zn kg-1, respectively. Rock phosphate was the only treatment that reduced GI-available Pb in both gastric and intestinal solutions, 23 and 92%, respectively. Alkaline organic treatments (LSB, NV) decreases Cd transmission through the food chain pathway, whereas rock phosphate decreases risk from exposure to Pb via the soil ingestion pathway. Alkaline organic treatments can reduce human exposure to Cd and Pb by reducing Zn phytotoxicity and revegetation of contaminated sites.

Abbreviations: LSB, lime-stabilized biosolid • RP, rock phosphate • NV, N-Viro Soil • GI, gastrointestinal


   INTRODUCTION
TOP
 NOTES
 ABSTRACT
 INTRODUCTION
METHODS AND MATERIALS
 RESULTS AND DISCUSSION
 SUMMARY AND CONCLUSIONS
 REFERENCES
 
MINING AND SMELTING of Pb and Zn metal ores are important sources of environmental degradation to soil and water resources (Dudka and Adriano, 1997; Lambert et al., 1997). Adverse environmental impacts from exposure to Cd, Pb, and Zn from contaminated smelter sites include risks to human health, phytotoxicity, contamination of water and soil, soil erosion, and ecotoxicity (Adriano et al., 1997; Pierzynski, 1997). Contaminated soil often presents an unacceptable risk to human and ecological health and must be remediated. Commonly used cleanup methods involved excavation and landfilling of smelter-contaminated soil. However, more permanent and less expensive in situ solutions have been favored in the last decade (Iskandar and Adriano, 1997).

Chemical immobilization is an in situ remediation method where inexpensive materials (e.g., fertilizer, waste products) are added to contaminated soil to reduce the solubility of heavy metal contaminants. Because contaminant solubility is related to its mobility and bioavailability, chemical immobilization may reduce environmental risk. Many studies have been conducted in the last decade using chemical amendments (including organic matter, alkaline material, and phosphate fertilizer for chemical remediation of Pb, Cd, and/or Zn in contaminated soil). Organic amendments used to immobilize Pb, Cd, and Zn in contaminated soil include municipal biosolids (sewage sludge), composts, manures, and peat (Brown et al., 1996; Pierzynski and Schwab, 1993). Alkaline materials have been used to immobilize metals in contaminated soils (Hooda and Alloway, 1996; Mench et al., 1994; Sappin-Didier et al., 1997). Phosphate amendments that can immobilize Pb in contaminated soils include hydroxyapatite (Boisson et al., 1999; Laperche et al., 1997; Ma et al., 1995) and soluble phosphates (McGowen, 2000; McGowen et al., 2001; Mench et al., 1994; Pierzynski and Schwab, 1993).

A few studies have investigated the ability of phosphate treatments to immobilize Cd and Zn in addition to Pb in smelter-contaminated soils (Hettiarachichi et al., 1998; Lambert et al., 1997; McGowen, 2000; McGowen et al., 2001). The success of chemical immobilization can be evaluated by its ability to reduce contaminant bioavailability and human exposure to heavy metal contaminants in treated, contaminated soil. Two important human exposure pathways are exposure to Cd through the food chain via plant uptake and exposure to Pb through ingestion of contaminated soil (Chaney and Ryan, 1994). Transmission of Cd, Pb, and Zn through the food chain is affected by the soil–plant barrier (Chaney and Giordano, 1977). The barrier limits transmission of metal through the food chain either by soil chemical processes that limit solubility or by plant senescence from phytotoxicity. The soil barrier limits Pb but not Cd or Zn uptake by plants (Chaney and Giordano, 1977). Cadmium can be transmitted through the food chain in levels that present risk to consumers (Chaney and Ryan, 1994). Although Zn is not a human health concern, phytotoxic levels of Zn can result in soil erosion by wind and water, thereby increasing human exposure to other metal contaminants (Pb, Cd). Lettuce (Lactuca sativa L.) has been used to assess heavy metal bioavailability in treated contaminated soil and food-chain risk to humans (Brown et al., 1996; Chaney and Ryan, 1994; Logan et al., 1997).

Incidental soil ingestion due to hand-to-mouth activity represents a significant direct exposure pathway to nondietary sources of heavy metals in contaminated areas (Chaney and Ryan, 1994; Duggan et al., 1985; Wixson and Davies, 1994). Animal models, as human surrogates, have been used to determine contaminant bioavailability via the ingestion pathway. Immature swine (Sus scrofa), rat (Rattus rattus), and rabbit (Oryctologus cuniculus) models have been used to simulate ingested heavy metal gastrointestinal (GI) bioavailability to humans (Casteel et al., 1996; Dieter et al., 1993; Rodriguez et al., 1999; Ruby et al., 1993, 1999). Determining bioavailability using animal models is expensive and requires specialized facilities and highly specialized personnel. Chemical in vitro laboratory methods that do not have the disadvantages associated with bioassay methods have been used to estimate heavy metal GI availability. The Physiologically Based Extraction Test (PBET), an in vitro chemical method that simulates the GI environment, has been shown to correlate well with animal feeding studies used to estimate Pb GI availability of ingested soil (Ruby et al., 1996, 1999). Use of the PBET method to evaluate chemical immobilization treatments of soil contaminated with Pb has been reported (Berti and Cunningham, 1997).

Long-term effectiveness and permanence are important criteria used to evaluate alternate remediation technologies (LaFornara, 1991). To be a successful remediation method, reductions in contaminant solubility and bioavailability demonstrated by chemical immobilization must also be long-term. Chemical immobilization of Pb in contaminated soil using phosphate treatments to produce Pb pyromorphite, a geochemically stable form of Pb in soil, may meet the permanence or long-term stability requirement for site remediation. However, identifying or quantifying many amorphous chemical immobilization products by spectroscopic analysis or other methods is not possible. An alternate approach to evaluate the stability of chemical immobilization products may be to determine the ability of heavy metals to remain insoluble (not extractable) upon acidification. Soil pH is one of the most important soil chemical properties affecting solubility of Cd, Zn, and Pb. Soil weathering often involves soil acidification, and most chemical immobilization reactions are pH dependent. Alkaline amendments reduce the concentration of heavy metals in soil solution by raising soil pH, thereby allowing the formation of insoluble metal precipitates, complexes, and secondary minerals (Mench et al., 1994; Pierzynski and Schwab, 1993). The ability of phosphate amendments to immobilize Pb through sorption, precipitation, and coprecipitation processes is also pH dependent (Chen et al., 1997; Xu et al., 1994; Zhang et al., 1997).

A successful chemical immobilization treatment of contaminated soil should provide long-term reductions in Cd and Zn phytoavailability and in Pb GI availability. Many studies have reported the ability of chemical immobilization treatments to reduce phytoavailability, solubility, or GI availability of Cd, Zn, and/or Pb. To our knowledge, all requirements of a successful chemical immobilization, as defined above, have not been evaluated. The use of soil acidification to provide information on the long-term stability of chemical immobilization products has not been reported. The objective of this work was to evaluate the effectiveness of organic, alkaline, and phosphate chemical immobilization amendments to reduce contaminant bioavailability, and to evaluate the effect of soil acidification on the extractability of metals in treated smelter-contaminated soils.


   METHODS AND MATERIALS
TOP
 NOTES
 ABSTRACT
 INTRODUCTION
 METHODS AND MATERIALS
RESULTS AND DISCUSSION
 SUMMARY AND CONCLUSIONS
 REFERENCES
 
Soil and Amendment Characteristics
The soils used in this experiment were collected from three sites contaminated by Zn and Pb milling and smelting operations in Oklahoma [Bartlesville (B4), Blackwell (BW), and Henryetta (H12)]. Soils were air-dried, sieved (<2 mm), and physical and chemical properties were determined (Table 1). Textural analysis was performed using the pipet method (Gee and Bauder, 1986). Soil pH was determined in 1:2 soil/0.01 M CaCl2 (Thomas, 1996), and electrical conductivity (EC) was determined in 1:2 soil/water (Rhoades, 1996). A 25-g subsample of soil was ground (corundum ball mill) to determine organic C by dry combustion using a Carlo-Erba NA 1500 (Nelson and Sommers, 1996).


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  		Table 1. Total metal content and properties of soils and amendments.

 
Total metal content of soils was estimated using USEPA Method 3050B (USEPA, 1986a). Toxicity Characteristic Leaching Procedure (TCLP) extraction was also performed on soil samples using a modified USEPA Method 1311 (USEPA, 1986b). Soil (5 g) was extracted with TCLP solution (0.1 M sodium acetate, pH 5.0) in a 125-mL polyethylene bottle on a reciprocal shaker for 18 h. Reagent-grade chemicals, trace-metal grade acids, and distilled–deionized water (>2 M{Omega}) were used in this study. Metal analyses were conducted using a Thermo Jarrell Ash IRIS Inductively Coupled Plasma–Atomic Emission Spectrophotometer (ICP).

Four chemical immobilization amendments were examined: a lime-stabilized municipal biosolid (LSB), a municipal biosolid–alkaline admixture blend marketed as N-Viro Soil (NV), North Carolina rock phosphate (RP), and an anaerobically digested municipal biosolid (AB). Both the LSB and the NV amendments were alkaline and had significant lime value (Table 1). Physical and chemical characteristics of the amendments (Table 1) were determined as described above. Each amendment was thoroughly incorporated into soil (100 g kg-1 soil) in plastic tubs. The amount of amendment applied was determined from a preliminary study where amendments were added at 10, 30, 100, and 300 g kg-1 to the Blackwell contaminated soil. The lowest application that resulted in lettuce growth for some treatments was 100 g kg-1. Application of apatite to Pb-contaminated soil at P/Pb molar ratio of 3:5 has been suggested to form chloropyromorphite [Pb5(PO4)3Cl] (Laperche et al., 1996; Ma et al.,1993; Zhang and Ryan, 1999; Zhang et al., 1998). Although molar ratio can be used as a basis for apatite treatment of Pb, applications based on the stochiometry of end products were not used because the Cd and Zn products formed after treatment of contaminated soils with RP are uncertain and immobilization products from LSB, NV, and AB are uncertain. Also, most insoluble heavy metal contaminants may not react with the amendment. Previous work has shown the amount of diammonium phosphate (DAP) required to reduce soluble Cd, Pb, and Zn in similar smelter-contaminated soils was <3:5 of the molar ratio of P/(Cd + Zn + Pb) suggesting some insoluble forms of these metals were not able to react with DAP (McGowen, 2000; McGowen et al., 2001). The 100 g kg-1 rock phosphate application corresponded to P/(Cd + Zn + Pb) molar ratios of 3.0 for the Henryetta soil, 2.0 for the Blackwell soil, and 1.3 for the Bartlesville soil (all disturbed soils).

All soil treatments were performed in triplicate. Soil moisture was adjusted to field capacity [0.033 MPa (0.33 bar), ca. 25% water], and the soils were incubated at 27°C for 90 d. Soil moisture was maintained, and the soils were thoroughly mixed at weekly intervals.

Contaminant Extractability and Bioavailability
The effect of soil treatments on contaminant extractability was evaluated by using the Potentially Bioavailable Assessment Sequential Extraction or PBASE method (Basta and Gradwohl, 2000). In this method, soil (1 g) was extracted sequentially with 20 mL of 0.5 M Ca(NO3)2 (E1), 1.0 M NaOAc (E2), 0.1 M Na2EDTA (E3), and 4 M HNO3 at 80°C (E4). Extracted Cd, Pb, and Zn were determined by ICP. Quality control procedures included use of blanks, spikes, standard reference materials, and comparison of summed PBASE fractions ({sum}E1–E4) for each soil treatment with total content (USEPA 3050) data. The summed PBASE fractions across all treatments and soils averaged 102.7% ± 8.7% of total Cd, 103.5% ± 8.2% of total Pb, and 100.3% ± 4.0% of total Zn.

To examine the effect of soil treatment on phytotoxicity and contaminant phytoavailability, lettuce (‘Paris Island Cos’) was grown in 15-cm pots containing 1-kg samples of soil over a 3-cm layer of vermiculite in a completely randomized design with three replicates. Excess salts were flushed from pots containing treated soil with distilled water until the EC was <0.5 dS m-1. Only very small amounts of Zn (<0.1% of total) were lost from flushing with water. Two weeks after emergence, lettuce plants were thinned to five in each pot. The lettuce bioassay was performed in a growth chamber under controlled conditions of 16 h light at 24°C and 8 h of darkness at 17°C for 90 d. The plants were watered as needed with a dilute water solution (1.0 g L-1) of Stern's Miracle-Gro plant food, which had a fertilizer analysis of 15–13–12 (N–P–K). Lettuce was harvested 2.5 cm above the soil surface and was rinsed with deionized water, then dried at 75°C for 48 h (Jones and Case, 1990). Lettuce samples were wet digested in hot nitric acid (Zarcinas et al., 1987).

The human GI availability of Pb from ingestion of treated soil was estimated using the Physiologically Based Extraction Test (PBET) (Ruby et al., 1996). The PBET method is a two-step sequential extraction of soil by a procedure that simulates GI biochemistry. The two extraction steps in PBET represent the gastric phase and the intestinal phase. In the first sequential extraction step, synthetic gastric solution (600 mL) and 1 mL of decanol (to prevent foaming) were added to a glass jar placed in a water bath (37°C). The gastric solution was prepared by adding 1.25 g pepsin (Sigma L-7000), 0.545 g citric acid monohydrate, 0.50 g malic acid, 420 µL lactic acid (Sigma L-1250), and 500 µL glacial acetic acid to a 1-L volumetric flask, diluting to volume with water, and adjusting the pH to 2.0 with HCl. Anaerobic conditions were maintained in the open reaction vessel throughout the experiment by argon sparging. Six grams of sieved soil (<250 µm) were placed in the reaction vessel, and the pH was adjusted to 2.0 with HCl while the solution was stirred with a paddle stirrer at 150 rpm. Solution pH was monitored at 5-min intervals and maintained at 2.0. After 1 h, three 10-mL aliquots of gastric solution were removed from the reaction vessel, filtered (0.45 µm), and stored at 4°C until analysis for gastric phase Pb by ICP. Fresh gastric solution was used to restore the reaction vessel volume. In the second sequential extraction step of PBET, the gastric reaction vessel solution was titrated to pH 7.0 by adding a length of dialysis tubing (cellulose ester, 100000 MWCO) containing NaHCO3 to the vessel (Medlin, 1997). After reaching pH 7.0 (ca. 90 min), 1.56 g of bile extract (B-8631, Sigma Chemical, St. Louis, MO) and 0.45 g of pancreatin (B-1750, Sigma Chemical) were added to the vessel and stirred for 1 h. Three 10-mL aliquots were removed and filtered (0.45 µm) for intestinal phase metal analysis. Solutions were acidified with HNO3 to pH < 2 and stored at 4°C until Pb analysis by ICP.

Acidification of Treated Soils
Soil/water (1:2) slurries of BW control and treated soils were acidified by adding predetermined amounts of 1 M HNO3 dropwise while paddle-stirring at 150 rpm to obtain target pH levels of 6, 5.5, and 4. The slurries were stirred continuously for 24 h and then placed in a forced-air oven at 70°C for 72 h. After drying, final pH of each soil was determined in 1:2 soil/0.01 M CaCl2. Acidification treatments were performed in triplicate. Acidified samples were extracted by the PBASE method to determine changes in Cd and Zn extractability, and by the PBET method to determine changes in Pb GI availability.


   RESULTS AND DISCUSSION
TOP
 NOTES
 ABSTRACT
 INTRODUCTION
 METHODS AND MATERIALS
 RESULTS AND DISCUSSION
SUMMARY AND CONCLUSIONS
 REFERENCES
 
Increases in soil pH of NV > LSB (Table 2) reflected the calcium carbonate equivalence of these materials (Table 1). Rock phosphate had little effect on soil pH, and AB decreased soil pH (Table 2). Similar decreases in soil pH were observed when biosolids from the same wastewater treatment plant were incubated with soil (Basta and Sloan, 1999). The high organic C content of the B4 and H12 soils compared with the BW soil did not result in increased pH buffer capacity to offset the lowering of soil pH from AB, because the organic C of these soils was due to coal residue at the contaminated sites rather than soil organic matter.


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  		Table 2. Effect of soil treatment on pH of smelter-contaminated soils.

 
Contaminant Extractability
The distribution of metal species within the four PBASE extraction fractions varied between metals (Fig. 1). Comparison of mean values of the three studied soils show 25% of soil Cd was in the E1 fraction and 77% was extracted by the first two extracts of the PBASE procedure (E1 and E2). The first two extracts averaged 38% of the total Pb for the three soils, of which only 0.5% was E1 extractable. Zinc showed intermediate extractability with mean values of 6 and 50% of Zn in E1 and E1 + E2 fractions, respectively. Because metal bioavailability is related to extractability, the relative bioavailability of the four metal fractions should be E1 > E2 > E3 > E4. Our data suggest metal solubility in smelter-contaminated soils is Cd > Zn > Pb, which agrees with earlier studies (Basta and Gradwohl, 2000; Elliott et al., 1986; Ma and Rao, 1997). Comparisons between soils show the most easily extracted E1 Cd was greater in the BW than B4 or H12 soils (Fig. 1). However, extractable metal in the E1 Pb or E1 Zn fraction was similar for all three soils. The E1 + E2 Pb and Zn fractions for the BW soil were greater than the B4 and H12 soils.



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  		Fig. 1. Distribution of Cd, Pb, and Zn in PBASE fractions E1–E4 expressed as a percentage of total metal content in Bartlesville (B4), Blackwell (BW), and Henryetta (H12) contaminated soils. Values at the top of graph bars represent total Cd, Pb, or Zn contents of the respective soil in mg kg-1.

 
The ability of soil treatments to reduce extractability of metal and reduce potential contaminant solubility and mobility was measured by the PBASE method. Recently, Cd in the E1 fraction was shown to be correlated with phytoavailable Cd, whereas Cd in the combined E1 + E2 fractions is dissolved by human GI fluids (Basta and Gradwohl, 2000). All treatments reduced easily extractable Cd in the E1 fraction (Fig. 2) in all soils, but the greatest reductions were associated with alkaline LSB and NV treatments. Increases for Cd in the E2 fraction and little change in E3 and E4 Cd fractions for alkaline treatments (LSB, NV) suggest that the Ca(NO3)2-extractable E1 Cd was converted to acid-labile E2-extractable forms (i.e., carbonates) (Fig. 2). These results are consistent with either adsorption of Cd on CaCO3 or precipitation of Cd as CdCO3. Soil treatment with RP decreased E1 Cd in the BW soil but had little effect on much smaller amounts of E1 Cd in the B4 or H12 soils. Hydroxyapatite has been shown to decrease 0.1 M Ca(NO3)2-extractable Cd in contaminated soil (Boisson et al., 1999). Both surface complexation and coprecipitation have been proposed as probable chemical mechanisms for reaction of soluble Cd with hydroxyapatite. However, greater reductions in the most soluble and easily extractable E1 Cd fraction were realized by alkaline LSB and NV not RP or AB treatments. These results suggest LSB and NV treatments should reduce Cd phytoavailability more than the RP or AB treatments.



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  		Fig. 2. Effect of immobilization treatments on Cd extractability in Blackwell (BW), Henryetta (H12), and Bartlesville (B4) contaminated soils as determined by the PBASE sequential extraction. LSB = lime-stabilized biosolids, NV = N-Viro Soil, RP = rock phosphate, and AB = anaerobically digested biosolids. Treatments with the same letter are not different at P < 0.05 within each soil–extract combination.

 
Similar to results for Cd, alkaline LSB and NV treatments resulted in the largest reductions in the E1 fraction of all three soils for Zn (Fig. 3). Similarly, smelter-contaminated soil treated with limestone reduced Zn extracted by 0.5 M KNO3 (Pierzynski and Schwab, 1993). Rock phosphate decreased Zn in the E1 fraction for the BW soil but had little effect on E1 Zn in H12 and B4 soils. However, treatment of a smelter-contaminated soil with hydroxyapatite at 0.5% (w/w) reduced 0.5 M Ca(NO3)2-extractable Zn from 201 to 102 mg kg-1 (Boisson et al., 1999). Surface complexation and coprecipitation have been proposed as probable chemical mechanisms for reaction of soluble Zn with hydroxyapatite. The smelter-contaminated soils in this study had much greater amounts of Zn extracted by Ca(NO3)2 solution (Table 1) than reported by Boisson et al. (1999). Perhaps a much larger application of rock phosphate (>10% w/w) is necessary to achieve further reduction of Zn in the E1 fraction. Treatment of smelter-contaminated soil with soluble phosphates, such as potassium phosphate (Pierzynski and Schwab, 1993) or diammonium phosphate (McGowen, 2000), greatly reduced Zn in the E1 fraction by forming Zn phosphate precipitates. Apparently, the low solubility of rock phosphate limited its ability to reduce soluble Zn by formation of Zn phosphate precipitates.



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  		Fig. 3. Effect of immobilization treatments on Zn extractability in Blackwell (BW), Henryetta (H12), and Bartlesville (B4) contaminated soils as determined by the PBASE sequential extraction. Immobilization treatment abbreviations are the same as listed in Fig. 2. Zn x 5 designation indicates the E1 values have been multiplied by 5. Treatments with the same letter are not different at P < 0.05 within each soil–extract combination.

 
Treatment of contaminated soil with nonalkaline sewage biosolids, AB, had no effect on the E1 Zn fraction for BW or B4 soils, and increased E1 Zn in the H12 soil. However, nonalkaline organic amendments (including poultry litter and cattle manure) reduced Zn extracted by 0.5 M KNO3 (Pierzynski and Schwab, 1993) apparently by chelation of Zn by organic solids. The inability of the AB treatment to reduce Zn in the E1 fraction can be attributed to the high Zn of 1675 mg kg-1 in the sewage biosolids (Table 1), and acidification associated with incubation of the AB treatment (Table 2). Previous studies with this biosolid applied to uncontaminated soil increased Zn in the E1 fraction (Basta and Sloan, 1999). The Zn content of the AB amendment lies between the 50th percentile (1202 mg kg-1) and 67th percentile (2756 mg kg-1) Zn content reported in the National Sewage Sludge Survey (USEPA, 1988). Thus, the AB amendment contained typical Zn levels commonly found in municipal biosolids. Apparently, organic amendments with low Zn are necessary to reduce potentially phytotoxic Zn in the E1 fraction.

Alkaline NV and LSB treatments reduced Pb extracted by E1 in the BW and H12 soils, but very little Pb (<20 mg kg-1) was in the E1 fraction (Fig. 4). The treatments with the higher organic C contents (LSB, AB) showed reductions in the E2 Pb fraction and concurrent increases in the E3 Pb fractions for treated BW soil. Blackwell soil treated with RP also showed reductions in the E2 Pb fraction and concurrent increases in the E3 Pb fraction. These results suggest Pb extractability may be reduced by sewage biosolids and by phosphate addition. Similarly, rock phosphate reduced Pb extractability as measured by the sequential extraction procedure of Tessier et al. (1979) in eight Pb-contaminated soils (Ma and Rao, 1997). Rock phosphate converted water soluble, exchangeable, carbonate, Fe–Mn oxide, and organic fractions into the most refractory (HF–HCl/HNO3) fraction apparently through phosphate rock dissolution and precipitation of fluoropyromorphite-like minerals. Water-soluble phosphate treatments of Pb-contaminated soil did not increase the residual fraction as measured by extraction with 4 M HNO3 (Pierzynski and Schwab, 1993). However, the Pb content of the soil studied by Pierzynski and Schwab (1993) (110 mg kg-1) was much lower than the Pb content of soil studied by Ma and Rao (1997) (ranging from 705 to 40100 mg kg-1) and the Pb content of soils of this study (ranging from 497 to 2650 mg kg-1). Perhaps high levels of Pb contamination are necessary to detect an increase of Pb in residual fractions due to phosphate addition.



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  		Fig. 4. Effect of immobilization treatments on Pb extractability in Blackwell (BW), Henryetta (H12), and Bartlesville (B4) contaminated soils as determined by the PBASE sequential extraction. Immobilization treatment abbreviations are the same as listed in Fig. 2. Pb x 100 designation indicates the E1 values have been multiplied by 100. Treatments with the same letter are not different at P < 0.05 within each soil–extract combination.

 
Contaminant Phytoavailability
Alkaline amendments (LSB and NV) significantly reduced the concentration of Cd and Zn in lettuce grown in the B4 soil compared with the control soil, whereas RP and AB did not (Fig. 5). Reductions in Cd phytoavailability in LSB and NV were consistent with reductions in E1 extractable Cd (Fig. 2) for the BW and B4 soils. The E1 fraction is highly correlated with phytoavailability of Cd in lettuce (Basta and Gradwohl, 2000). Control and AB data for the BW soil are not presented in Fig. 5 because lettuce did not grow on these soils due to Zn phytotoxicity. Phytotoxic levels of available Zn associated with the E1 fraction of 1188 mg Zn kg-1 soil in the BW soil were significantly reduced with the addition of the LS, NV, and RP amendments to 166, 25, and 784 mg Zn kg-1, respectively (Fig. 3). The extractable E1 Zn of Control, AB > RP > LS, NV mirrors the growth of lettuce on the BW soil (Fig. 6). In general, other treatments had little effect on Cd or Zn in the H12 soil. The AB treatment increased the Zn content of lettuce grown in treated H12 soil (Fig. 5). Addition of nonalkaline biosolids (AB) increased E1 extractable Zn for the B4 and H12 soils, and plant uptake of Zn from these soils. Pierzynski and Schwab (1993) showed that the addition of organic amendments (i.e., cattle or poultry manure) reduced soybean [Glycine max (L.) Merr.] tissue Zn concentrations grown in metal-contaminated soil. Increases in plant-available Zn in AB-amended soil in this study was due to high levels of bioavailable Zn added with the amendment (Basta and Sloan, 1999).



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  		Fig. 5. The effect of immobilization treatments on Cd, Pb, and Zn phytoavailability of lettuce grown on Blackwell (BW), Henryetta (H12), and Bartlesville (B4) contaminated soils. Immobilization treatment abbreviations are the same as listed in Fig. 2. Treatments with the same letter are not different at P < 0.05 within each soil.

 


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  		Fig. 6. Lettuce grown in control and treated Blackwell contaminated soil. Photo taken at time of harvest. Immobilization treatment abbreviations are the same as listed in Fig. 2.

 
Lettuce does not readily translocate Pb to aboveground tissue as compared with Cd or Zn. Some differences in lettuce Pb concentrations were found (Fig. 5), but all lettuce Pb concentrations were low (<6 mg kg-1). Consequently, lettuce Pb concentrations were very low, making changes due to treatment effects difficult to measure.

Gastrointestinal Availability of Lead
The greatest reduction of Pb in the gastric phase of 23% was obtained with the RP-treated BW soil (P < 0.10) (Fig. 7). The intestinal phase concentrations of Pb in soils were lower than their respective gastric concentrations (Fig. 7). Several treatments reduced available Pb in the intestinal phase. N-Viro, RP, and AB amendments reduced intestinal Pb by 55, 92, and 58%, respectively. Although the intestinal system is responsible for most absorption processes, Ruby et al. (1996) preferred gastric phase Pb as a measure of Pb bioaccessibility based on the correlation of their in vivo (Sprague-Dawley rat) and in vitro results (r2 of 0.93 for gastric and r2 of 0.76 for intestinal, n = 7). They attributed the lower correlation of the intestinal phase data to poor reproducibility of their intestinal simulation. Because the acidic (pH 2.0) gastric phase solution is more aggressive at dissolving Pb than the neutral (pH 7.0) intestinal phase solution, Pb extracted from treated soils from the gastric phase conditions of the PBET method offers a more conservative estimate of GI bioavailability than Pb extracted from the intestinal phase. Assuming gastric metal concentrations offer a worst-case scenario, a small reduction in bioavailable Pb was observed with the RP amendment in the BW soil (P < 0.10). Of all four treatments, RP showed the largest reduction (>92%) of intestinal Pb.



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  		Fig. 7. Effect of immobilization treatments on gastrointestinal Pb in the Blackwell soil as measured using the PBET method. Immobilization treatment abbreviations are the same as listed in Fig. 2. Treatments with the same letter are not different at P < 0.05 within gastric or intestinal extracts.

 
Acidification of Treated Soils
Acidification of BW Control soil resulted in a redistribution of Zn and Cd from the E2, E3, and E4 fractions into the E1 fraction (Fig. 8). The ability of LSB, NV, and AB soil treatments to reduce Zn and Cd in the readily extractable E1 fraction was lost when the BW soil was acidified to pH < 6. Soil acidification increased E1 extractable Cd and Zn in RP-treated BW soil, but this effect was smaller than other treatments. However, the E1 Zn levels of >3000 mg kg-1 at pH < 6 are phytotoxic. Therefore, acidification that results in pH < 6 for the BW soil results in phytotoxicity regardless of soil treatment.



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  		Fig. 8. The effect of soil acidification on Cd and Zn in the E1 PBASE fraction and gastrointestinal Pb, measured in the gastric phase of the PBET method, for the Blackwell contaminated soil. Immobilization treatment abbreviations are the same as listed in Fig. 2.

 
Acidification increased the GI available Pb, measured by the gastric phase of the PBET method, for biosolid treatment (Fig. 8). Acidification of the RP-treated BW soil decreased GI available Pb. Decreased PBET Pb is likely due to dissolution of apatite under acidic conditions with subsequent formation of extremely stable Pb pyromorphites (Chen et al., 1997; Laperche et al., 1996; Ma et al., 1994, 1995; Ruby et al., 1994; Zhang et al., 1997) as well as adsorption and coprecipitation of Pb (Takeuchi and Arai, 1990).


   SUMMARY AND CONCLUSIONS
TOP
 NOTES
 ABSTRACT
 INTRODUCTION
 METHODS AND MATERIALS
 RESULTS AND DISCUSSION
 SUMMARY AND CONCLUSIONS
REFERENCES
 
The largest reductions in metal extractability and phytoavailability were realized using alkaline organic treatments (LSB and NV). However, the products of the LSB and NV treatments (i.e., carbonates, chelates) were not stable under acidic conditions of the acidic pH 2.0 gastric solution of the PBET procedure. Reductions in GI bioavailability measured in the neutral pH 7.0 intestinal solution of the PBET procedure were found for NV. However, the more aggressive acidic gastric phase has been recommended to provide conservative estimates of GI bioavailability (Ruby et al., 1999). The ability of LSB and NV soil treatments to reduce Zn and Cd extractability was lost when soil was acidified to pH < 6. Large applications of LSB or NV that prevent development of acidic soil conditions (pH < 6) are necessary to maintain reductions in metal extractability and phytoavailability and to prevent Zn phytotoxicity.

Treatment of contaminated soil with AB increased phytoavailable Zn and did not reduce plant uptake of Cd or Pb. In part, increases in phytoavailable Zn were due to high levels of Zn considered typical to biosolids. Use of organic amendments with low Zn content will likely decrease phytoavailable Zn. Biosolid immobilization products were not stable under acidification (pH < 6), and did not decrease GI available Pb in the gastric phase.

Rock phosphate decreased metal extractability only in the BW soil and had little effect on metal phytoavailability in the contaminated soils. The reductions in metal extractability obtained with RP in the BW soil were much smaller than with the alkaline LSB and NV treatments. Consequently, reductions in plant uptake of metal were LSB, NV > RP. However, the RP immobilization products (i.e., metal pyromorphites) were more stable under acidification (pH < 5) and, unlike other treatments, reductions in GI availability and extractability of Pb were realized under acidic soil conditions. Treatment with rock phosphate would reduce Pb bioavailability associated with ingestion of contaminated soils by humans. Some of the RP-treatment products should be stable under conditions that favor soil acidification, and repeat applications of RP would not be necessary upon soil acidification.

None of the soil treatments meet all the criteria used in this study to determine successful remediation of Cd-, Pb-, and Zn-contaminated soil. In general, alkaline organic treatments reduce metal extractability and phytoavailability but not GI availability. Rock phosphate treatment reduces metal extractability and GI availability of Pb, but may not reduce Zn and Cd phytoavailability. Transmission of Cd through the food chain via plant uptake and exposure to Pb through ingestion of contaminated soil are important human exposure pathways. Use of alkaline organic treatments (LSB, NV) are well suited to decrease Cd transmission through the food chain pathway, whereas rock phosphate is well suited to decrease exposure to Pb via the soil ingestion pathway. Alkaline organic treatments can reduce human exposure to Cd and Pb from eroded soil by reducing Zn phytotoxicity and by revegetating contaminated sites.


   ACKNOWLEDGMENTS
 
The authors acknowledge the assistance of Scott Thompson and others at the Oklahoma Department of Environmental Quality in obtaining materials and information. The authors thank associate editor Dr. David R. Parker, the anonymous reviewers for their comments that led to an improved manuscript, and Susan Bennett Basta for copyediting the manuscript.


   NOTES
TOP
 NOTES
ABSTRACT
 INTRODUCTION
 METHODS AND MATERIALS
 RESULTS AND DISCUSSION
 SUMMARY AND CONCLUSIONS
 REFERENCES
 
Published with approval of the Director, Oklahoma Agric. Exp. Stn.


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




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