Calcium–phosphate treatment of contaminated soil for arsenic immobilization

Applied Geochemistry 28 (2013) 145–154

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Calcium–phosphate treatment of contaminated soil for arsenic immobilization Ghanashyam Neupane a,b,⇑, Rona J. Donahoe a a b

Department of Geological Sciences, University of Alabama, Box 870338, Tuscaloosa, AL 35487, United States University of Idaho-Idaho Falls, 1776 Science Center Drive, Suite 306, Idaho Falls, ID 83402, United States

a r t i c l e

i n f o

Article history: Received 26 December 2011 Accepted 12 October 2012 Available online 26 October 2012 Editorial handling by J. Routh

a b s t r a c t The application of As-based herbicides at several industrial sites has resulted in numerous localized areas of As-contaminated soil. In this study, an As-contaminated soil (As = 278 mg/kg) collected from an industrial site located in the southeastern USA was subjected to inorganic phosphate (Pi) treatments. Although Pi treatments have been previously used for flushing As from contaminated soils, in this study, contaminated soil was amended with Pi to study the possible immobilization of As through a co-precipitation mechanism. Specifically, the Pi amendment was aimed at simultaneous flushing of As from the soil with orthophosphoric acid and co-precipitating it as Ca–phosphate–arsenate phases. Bench-scale Pi treatment experiments were performed at different pH conditions, with and without the addition of Ca. Sorption of Pi on BH soil in the presence or absence of additional Ca was determined, along with the associated mobilization of As from the soil. A significant amount of the HNO3-digestible As (up to 55% at pH 4, 10–15% at pH 8, and 30% at pH 11) was released from the contaminated soil during the Pi sorption experiments. This increased mobility of As after the addition of Pi resulted from the competitive desorption of As from the soil. Although Pi sorption at high pH (>8) was largely controlled by precipitation, As did not co-precipitate with Pi. Aqueous geochemical modeling indicated that the lack of As co-precipitation during Pi-only treatment primarily resulted from the deficiency of Ca in the system. When additional Ca (16.9 mmol) was supplied along with Pi (3.38 mmol), the mobility of As decreased significantly at circum-neutral to high solution pH. Geochemical modeling suggested that the leachable As in the soil was potentially precipitated as As-bearing Ca–Pi phases. X-ray diffraction analysis of precipitates separated from the treated soil and from the synthetic leachate confirmed that the formation of a poorly crystalline carbonate apatite phase occurred as a consequence of the treatment. The results of this study support the potential application of Ca–Pi treatment for remediation of As-contaminated soil at environmentally relevant pH conditions. Ó 2012 Elsevier Ltd. All rights reserved.

1. Introduction Arsenic is a toxic element, an acute dose of 50–300 mg As being considered to be lethal for humans due to gastrointestinal, respiratory, cardiovascular, neurological or other body system failures (ATSDR, 2000). Similarly, chronic ingestion of As, either through food or water, is responsible for several types of cancer (Jackson and Grainge, 1975; Bates et al., 1992; Karagas et al., 2002). Elevated concentrations of As in the environment can occur by natural processes as well as anthropogenic activities (Smedley and Kinniburgh, 2002). Groundwater As-enrichment from natural sources has adversely affected the health of millions of people in several areas of the world (Nordstrom, 2002; Hossain, 2006; Bundschuh et al., 2012). Arsenic contamination has occurred through the use of As-based pesticides (Simcox et al., 1995; Robinson et al., 2007), herbicides (Yang and Donahoe, 2007) and wood preservatives (Morrell et al., 2003), and through mining and smelting activities (Carbonell-Barrachina et al., 2004). ⇑ Corresponding author. Tel.: +1 208 282 7842; fax: +1 208 282 7929. E-mail address: [email protected] (G. Neupane). 0883-2927/$ - see front matter Ó 2012 Elsevier Ltd. All rights reserved. http://dx.doi.org/10.1016/j.apgeochem.2012.10.011

In the USA, As-based compounds have been used for several purposes such as agricultural pesticides and herbicides, wood preservatives and glass production (Welch et al., 2000). Although the widespread use of different arsenical compounds gradually declined after the 1960s when the adverse environmental effects of As were realized, their legacy in environmental pollution is still pronounced in many areas of the USA (Welch et al., 2000). Several industrial sites in the southeastern USA which received intensive applications of arsenical herbicides during the 1950s continue to leach As into surface water and groundwater and pose considerable danger (Yang and Donahoe, 2007; Qi and Donahoe, 2008). Regulatory agencies such as the United States Environmental Protection Agency (USEPA) have developed, and are enforcing more stringent standards for As levels in drinking water. For example, the USEPA lowered the maximum contaminant level (MCL) for As in drinking water from 50 lg/L to 10 lg/L in 2006. This required the development and implementation of As remediation strategies at several contaminated sites to avoid potential contamination of local potable water sources. Several techniques such as isolation and separation (Mulligan et al., 2001), solidification and stabilization (Yang et al., 2007),

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vitrification (Mulligan et al., 2001), soil flushing (Alam et al., 2001), electrokinetic treatment (Yuan and Chiang, 2007), and bioremediation (Ma et al., 2001) have been proposed for remediation of As-contaminated soils. Previous bench-scale As-remediation experiments performed in the authors’ laboratory (Yang et al., 2007; Bhattacharyya et al., 2009; Neupane et al., 2010) have successfully demonstrated the in situ chemical fixation of As in contaminated soil and coal fly ash with ferrous sulfate solutions. Flushing of As from contaminated soil and soil constituents with solutions containing competing ligands such as inorganic phosphate (Pi) and OH has been studied extensively (Johnson and Barnard, 1979; Peryea and Kammereck, 1997; Wasay et al., 2000; Alam et al., 2001; Tokunaga and Hakuta, 2002; Kaplan and Knox, 2004). Because of its similar chemical characteristics, Pi has a high potential to desorb As(V) from soil through ion exchange reactions (Woolson et al., 1973; Jackson and Miller, 2000; Wasay et al., 2000; Alam et al., 2001; Kaplan and Knox, 2004). In particular, phosphoric acid is very effective in rapidly displacing As from contaminated soil by providing Pi to compete with As(V) for adsorption sites, and by dissolving As-rich metal oxides present in soils (Tokunaga and Hakuta, 2002). A few studies (e.g., Grisafe and Hummel, 1970; Twidwell et al., 1994) have also demonstrated co-precipitation of As with Pi minerals. In these studies, As(V) substituted for Pi to form continuous solid solutions of fluor- and chlor-apatites (Grisafe and Hummel, 1970). Subsequently, Mahapatra et al. (1987) successfully synthesized arsenate hydroxyapatite. Similarly, Twidwell et al. (1994) and others (Wilson, 1998; Orser, 2001) indicated removal of As from solution by precipitation of Ca–Pi–As(V) minerals at pH 8–10. Their study indicated that Pi–As(V) hydroxyapatite with a P:As ratio P7 has a low solubility (<10 lg/L) and greater stability against atmospheric CO2 compared to other Pi–As(V) apatites. In this study, an As-contaminated soil collected from an industrial site located in the southeastern USA was subjected to a Ca–Pi treatment. The objective of this study is to test the potential application of phosphate-based treatment for remediation of As-contaminated soils. The remediation method evaluated in this study aims at flushing As from contaminated soil with phosphoric acid, and immobilizing the same through precipitation as Ca–Pi–As(V) minerals. 2. Materials and methods 2.1. Contaminated site and soil sample collection An As-contaminated soil sample was collected from an industrial site (BH) located near the Gulf Coast in the southeastern USA. The soil had developed from Quaternary undifferentiated marine and fluvial sediments. With increasing depth, four subsurface units: an unconfined sandy aquifer, a silty peat semiconfining bed, a semi-confined sand aquifer, and a confining silt/ clay layer, have been reported at the site (Yang and Donahoe, 2007). The regional climate of the contaminated area is humid and semi-tropical with an annual rainfall of about 162 cm (Yang and Donahoe, 2007). The rainy season at the sampling site begins in May and ends in November, and accounts for ca. 60% of the annual rainfall (Black, 1993). Temperature is moderate, with an average high of 28 °C in summer and an average low of 10 °C in winter (Yang and Donahoe, 2007). The soil at this site was contaminated during the 1950s through the one-time application of arsenic trioxide (arsenolite) rich herbicide produced at Anaconda smelter in Montana (Yang and Donahoe, 2007). Originally, arsenolite was applied as a 1-in. (2.5 cm) thick layer on top of the natural soil surface, and covered with a 5–10 cm thick layer of limestone gravel. During the last five decades, the arsenolite layer has weathered and dispersed into the

soil, creating a potential risk for As exposure to the local population (Yang and Donahoe, 2007). Because earlier studies (e.g., Yang and Donahoe, 2007) had indicated that As was mostly confined to the vadose zone, soil samples were collected at a single location from 10 to 60 cm below the surface and placed in two 19 L-buckets. The soil was air-dried, passed through a 2 mm sieve, homogenized, and stored at room temperature for chemical/mineralogical characterizations and Pi treatments. 2.2. Characterization of BH soil Arsenic and other elements in the BH soil samples were extracted by microwave-assisted HNO3 partial digestion (MWD) (USEPA, 2007), and determined with a Perkin Elmer Optima 3000DV inductively coupled plasma-optical emission spectrometer (ICP-OES). The soil pH was measured according to USEPA Method 9045D (USEPA, 2004). The single-point BET specific surface area of the soil sample was determined using ASTM D4567-03 method (ASTM, 2003). Minerals present in the bulk soil, the soil clay fraction, and experimental precipitates were determined by X-ray diffraction (XRD) using a Brüker D8 Advance powder diffractometer. The bulk soil sample was powdered in an iron ShatterboxÒ mill for 2 min for XRD analysis. The clay-sized soil fraction was separated by flotation according to Stoke’s Law. The XRD data were analyzed with DIFRACplus EVA version 11.0.03 (XRD data evaluation and presentation software provided by Bruker AXS), using the International Center for Diffraction Data, ICDD PDF-2 Release 2005 database. 2.3. Experiments 2.3.1. Mobilization of As from BH soil as a function of pH Mobilization of As from BH soil as a function pH was studied in batch experiments. For each experimental setup, 3.0 g of soil was weighed in a 50 mL centrifuge tube and then 45 mL of 0.1 mol NaCl solution was added. A series of similar experimental sets were prepared and agitated on an orbital platform shaker for 24 h. The pH of the soil mixture was set initially and adjusted periodically when needed in the range of 3–12 using 0.1 mol or 1.0 mol HCl and NaOH solutions. The supernatant solutions were filtered through 0.2 lm nylon syringe filters after 20 min centrifugation at 13,000g, acidified to 2% HNO3 with OPTIMAÒ ultrapure HNO3, and stored at 4 °C in a refrigerator until chemical analysis by ICPOES. The mobilization of As and Ca from soil was calculated using Eq. (1) and expressed as % mobilized.

% Mobilized ¼

Amount leached to solution  100% MWD extracted amount in soil

ð1Þ

2.3.2. Pi sorption and mobilization of As from BH soil The sorption of Pi on BH soil was evaluated by constructing sorption isotherms, and by conducting sorption experiments as a function of pH. The term ‘sorption’ is used to encompass all processes (adsorption, absorption, and precipitation) which transfer aqueous sorbate (e.g., Pi) to the solid phase (Sposito, 1986). Each sorption experiment was performed at a 1:15 solid:liquid ratio by mixing 3.0 g of soil with 45 mL of 0.1 mol NaCl electrolyte in a 50 mL centrifuge tube. The experiments were performed in duplicate, but each experiment was treated as an individual sample, rather than calculating an average value for the duplicate samples. The phosphate stock solution was prepared from 85% H3PO4; the Pi concentration was determined by ICP-OES analysis. Further dilutions of the Pi stock solution were performed with respect to its analyzed concentration. Isotherms for phosphate sorption on

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BH soil at pH 4, 8, and 11 were constructed using 0.15–13.24 mmol initial Pi concentrations. Sorption as a function of pH was determined for 0.57 mmol and 3.38 mmol initial Pi concentrations in the pH range from 3 to 12. The pH of the soil mixture was adjusted at the beginning of the experiment using HCl and NaOH solutions. Furthermore, pH was checked intermittently and adjusted, if necessary. The tubes were agitated for 24 h, and the samples were centrifuged, filtered, acidified and stored for chemical analysis. During ICP-OES analyses, quality control check standards were run after analysis of each batch of 10 samples. The relative standard deviations of As, Ca and P concentrations in the check standards were generally within ±5%, and never exceeded ±10%. The sorption of Pi on BH soil was determined by calculating the difference in concentrations of Pi in initial solution and equilibrium solution. Finally, the mobilization of As and Ca from soil during Pi sorption experiments was calculated using Eq. (1). 2.3.3. Sorption isotherm modeling Experimental sorption data were fitted with the Freundlich isotherm model (Eq. (2)):

qe ¼ KCne

ð2Þ

where qe is the amount of sorbate (Pi) sorbed on BH soil (mmol/kg), Ce is the equilibrium sorbate concentration (mmol), K (L/kg) and n (dimensionless) are constants. Freundlich isotherm model parameters were obtained by converting Eq. (1) into linear form (Eq. (3)):

log qe ¼ log K þ n log C

ð3Þ

where values of n and K can be determined from the slope and yintercept of a graph obtained by plotting log Ce along the x-axis and log qe along y-axis (Limousin et al., 2007). For each isotherm, the goodness of model fit was determined by calculating the coefficient of determination (R2) with:

R2 ¼ Pn

i¼1 ðqm

Pn 

i¼1 ðqm  e Þ2 þ q

 e Þ2 q Pn 2 i¼1 ðqm  qe Þ

ð4Þ

where qe is as defined in Eq. (2), qm is the model-derived maximum e is the average experimental adsorbed amount (mmol/kg), and q sorbed amount (mmol/kg). The closer the value of R2 is to 1.0, the better the model is in predicting sorption. The R2 value for each isotherm was obtained by using the experimental and corresponding model datasets. 2.3.4. Ca–Pi treatment of BH soil The contaminated soil was treated with Ca and Pi in batch experiments. For each experiment, 3.0 g of soil was weighed in a 50 mL centrifuge tube and 45 mL of a Ca–Pi treatment solution was added to it. The Ca treatment solutions were prepared from dry, analytical grade CaCl2, and the Ca concentrations were determined by ICP-OES analysis. Two sets of batch treatment experiments were performed over the pH range from 3 to 12. The first set of experiments was conducted with Ca and Pi concentrations of 5.63 mmol and 3.38 mmol, respectively, yielding a 1:67 initial Ca:Pi ratio in the soil mixture. The second set of experiments used the same concentration of Pi, but a much higher concentration of Ca (16.90 mmol), produced an initial Ca:Pi ratio of 5:1. The pH of the soil mixture was adjusted with HCl and NaOH solutions and agitated for 24 h. Finally, the tubes were centrifuged and the supernatant solutions were separated by filtration and stored in a refrigerator until chemical analysis. The precipitate in the treated soil was analyzed by X-ray diffraction to identify the mineral phases. The % mobilization of As during soil treatments was calculated using Eq. (1). However, Eq. (1) was modified to Eq. (5) so as to include the externally supplied Ca while calculating % mobilization of Ca during the treatment:

% Mobilized ¼

147

Ca in equilibrium solution  externally supplied Ca MWD extracted amount of Ca  100% ð5Þ

According to Eq. (5), 0% mobilization indicates neither release of Ca from soil to solution nor transfer of Ca from solution to soil. Therefore, a soil solution with 0% mobilization of Ca contains only the externally supplied aqueous Ca. Positive % mobilization values indicate that the equilibrium solution is enriched in Ca due to the release of Ca from soil, while negative % mobilization values mean that some of the externally supplied Ca is transferred to the soil solid phase by sorption, mostly by precipitation mechanism. During Ca–Pi treatments, 45 mL of 5.63 mmol or 16.9 mmol Ca, along with 3.38 mmol Pi, were reacted with 3 g of BH soil. Therefore, complete transfer of Ca from the treatment solution to the solid phase would result in calculated Ca mobilities of about 14% (for 5.63 mmol added Ca) and 42% (for 16.9 mmol added Ca), with respect to the soil’s 592 mmol/kg of HNO3-digestible Ca. For As, however, the calculated mobilization should be always P0%. A 0% mobilized value indicates that no As is released from soil to solution during the Ca–Pi treatment. 2.3.5. Precipitation of Ca–Pi–As(V) phases from synthetic leachate Calcium–Pi–As(V) solid phases were precipitated from synthetic leachate prepared by mixing Ca, Pi, and As(V) stock solutions in 0.1 mol NaCl. The calculated amount of vacuum desiccator dried Na2HAsO47H2O was used for preparation of the As(V) stock solution. In addition, the concentration of As in the stock solution was determined by ICP-OES analysis. The initial synthetic leachate with As(V) = 0.48 mmol, Pi = 3.38 mmol, and Ca = 6.44 mmol was stirred for 24 h in a 250 mL beaker, while in contact with air, at room temperature (ca. 21 °C) to precipitate Ca–Pi–As(V) solid phases. The As(V):Pi ratio was selected to be 1:7, as suggested by Wilson (1998), whereas the Ca:(As(V) + Pi) ratio was fixed at 1.67. The pH of the synthetic leachate was adjusted to 8 with NaOH and HCl solutions. At the end of the 24 h period, the precipitate formed was separated by vacuum filtration and analyzed by XRD. 2.4. PHREEQC simulations and potential precipitation of Ca–Pi–As(V) minerals The aqueous chemical data were used to calculate the saturation index (SI) of several Ca, Ca–As(V), Ca–Pi, and Ca–Pi–As(V) minerals with the PHREEQC geochemical code (Parkhurst and Appelo, 1999), using the llnl.dat database. Thermodynamic data for several potential minerals and aqueous species (see Supplementary Tables S1 and S2) were collected from additional sources (Smith and Martell, 1976; Bothe and Brown, 1999; Wilson, 1998; Montastruc et al., 2003) and incorporated into the llnl.dat database. Most of thermodynamic data for different solid phases compiled for this study were determined/estimated at 20–25 °C. The laboratory temperature during the experiments performed in this study was in the range of 20–22 °C, and the thermodynamic data were used as obtained without any temperature correction. During the PHREEQC simulations, all solutions were allowed to equilibrate with atmospheric CO2 and O2. The precipitation potential of a mineral was determined by evaluating its saturation state (undersaturation or oversaturation). The degree of undersaturation or oversaturation of an aqueous sample with respect to a particular mineral was determined in terms of the SI, calculated using the following equation:

SI ¼ logðIAP=K sp Þ

ð6Þ

where IAP is the ion activity product and Ksp is the solubility constant for a particular mineral (Drever, 1997). If the SI for a mineral

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is greater than 0, then the aqueous solution is considered to be oversaturated with respect to that particular mineral, and the solution is likely to spontaneously precipitate that mineral, depending on kinetic constraints. Speciation modeling was performed by PHREEQC for different treatment scenarios. The chemical data obtained as a function of pH from the batch experiments without external supply of Pi or Ca were used as primary chemical data. Initially, the amounts of As and Ca mobilized as a function of pH were evaluated for potential precipitation. Subsequently, the amounts of Pi and Ca supplied to the soil during treatments were added to the primary chemical data, and the potential for precipitation of different minerals from the augmented solutions was evaluated. 3. Results

isotherm model is able to describe the experimental data at pH 4 (R2 = 0.93) and 11 (R2 = 0.93). While fitting the experimental isotherm data for pH 8 with the Freundlich model, the data points with Ce > 6.5 mmol were not used due to an apparent change in sorption mechanism (i.e., precipitation). With this exception, the Freundlich model resulted in a very good fit for pH 8 experimental isotherm data (R2 = 0.97). The n and K (L/kg) values for Pi sorption on BH soil at pH 4, 8, and 11 are 0.36, 0.48, and 0.47, and 51.5, 4.6, and 12.2, respectively. Phosphate sorption on BH soil as a function of pH for two initial Pi concentrations (0.57 mmol and 3.38 mmol) is shown in Fig. 2, where the sorption of Pi on soil is presented as the percentage sorbed. Sorption of Pi is very high at acidic pH (<5) (except at very low pH for Ci = 3.38 mmol), low between pH 5.5–8, and moderately high at pH > 9.

3.1. Characterization of BH soil

3.3. Mobilization of As and Ca from BH soil

BH soil is a basic (pH  8.4), medium dark gray (N4), sandy-clay loam. This soil mostly consists of quartz, calcite, and kaolinite with minor amount of gehlenite, montmorillonite, and muscovite. The concentration of As in BH soil is 278 mg/kg (3.71 mmol/kg). The background concentration of As in soil near the sampling site has been reported to be <3 mg/kg (Yang and Donahoe, 2007). BH soil has high concentrations of Ca (23,700 mg/kg), Fe (1360 mg/kg), Al (5860 mg/kg), Mg (5600 mg/kg), K (765 mg/kg), and Na (280 mg/kg). The organic C content (loss-on-ignition) of this soil is 6.5 mg/g (Neupane et al., 2010) and the surface area of the bulk soil is 2.65 m2/g, as determined by the single point BET method. Speciation of As in BH soil was studied in detail by Yang and Donahoe (2007), who used sequential chemical extraction followed by electron microprobe analysis, scanning electron microscopy, l-XRD, and X-ray absorption spectroscopy (XAS). Yang and Donahoe (2007) reported that most of the As in BH soil remains sorbed to amorphous Al- and Fe-hydroxides. However, trace amounts of some As-rich minerals (e.g., phaunouxite) are also present in the contaminated soil (Yang and Donahoe, 2007). Sequential chemical extraction and l-XANES studies indicated that most of the original As(III) arsenolite herbicide applied more than 5 decades ago has been oxidized to As(V) by weathering processes (Yang and Donahoe, 2007).

3.3.1. As and Ca mobilization as a function of pH Data showing the mobilization of As and Ca from BH soil as a function of pH are presented in Fig. 3 in terms of mmol sorbate/ kg dry soil (Fig. 3a) and as % mobilized (Fig. 3b). Each set of duplicate batch experiments containing 3.0 g soil and 45 mL of 0.1 mol NaCl was adjusted to a particular pH level with 0.1 mol or 1.0 mol solutions of NaOH and HCl. Mobilization of As increased with increasing pH (with the exception of high mobilization at pH 3) from 4 to 12. However, the mobility of Ca decreased with increasing pH (Fig. 3a and b). The SI values (Fig. 3c) for several Ca minerals with or without As were calculated as a function of pH with the input solution compositions of non-Pi leachate solutions (Fig. 3a). These solutions were undersaturated with respect to portlandite and all Ca-As(V) minerals included in Table S2. Calcite was the only one mineral at saturation level at pH > 8 in non-Pi soil leachate solutions (Fig. 3c).

The experimental isotherms with Freundlich model fits are shown in Fig. 1. All three isotherms show a greater sorption of Pi at higher initial concentration, while relatively greater proportions of Pi are sorbed at lower initial concentration. The Freundlich

3.3.2. As and Ca mobilization as a function of Pi During Pi sorption isotherm experiments (Section 3.2), varying amounts of As and Ca were released from BH soil. Fig. 4a and b show the mobilization of As and Ca as a function of Pi at pH 4, 8 and 11, in terms of the % mobilized (Eq. (1)). Both elements showed greater mobility at pH 4 than at pH 8 and 11. Mobilization of Ca at pH 4 was two orders of magnitude greater than at pH 8 and 11. The mobility of Ca decreased with increasing pH (pH 4 > pH 8 > pH 11). The mobilities of As and Ca increased with increasing concentrations of Pi at pH 4. The mobility of Ca leveled off at ca. 42–45% mobilization for Ci > 4 mmol Pi (at pH 4), whereas the rate of increase in As mobility slowed for Ci > 6 mmol Pi. Although the

Fig. 1. Phosphate sorption on BH soil in 0.1 mol NaCl at pH 4, 8, and 11. Lines represent Freundlich isotherm model fits to data.

Fig. 2. Phosphate sorption on BH soil as a function of pH for 0.57 mmol or 3.38 mmol initial Pi concentrations in 0.1 mol NaCl.

3.2. Pi sorption on BH soil

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149

Fig. 4. Percent mobilized (a) As and (b) Ca as a function of the initial concentration of Pi and calculated with respect to the HNO3-digestible concentrations of these elements in BH soil (As = 3.71 mmol/kg; Ca = 592 mmol/kg). The inset in (b) shows the expanded y-axis for pH 8 and 11.

Fig. 3. Mobilized As and Ca as a function of pH during batch leaching of 3 g soil in 45 mL 0.1 mol NaCl. (a) Mobilized As and Ca given in mmol/kg dry soil; (b) percent mobilized calculated with respect to HNO3-digestible soil concentrations (As = 3.71 mmol/kg; Ca = 592 mmol/kg); and (c) mineral SI calculated using PHREEQC. In (c) – 1: Ca3(AsO4)2, 2: Ca3(AsO4)23.66H2O, 3: Ca3(AsO4)24.25H2O, 4: Ca4(OH)2(AsO4)24H2O, 5: Ca5(AsO4)3OH, 6: CaHAsO4H2O, 7: calcite, 8: Ferrarsite, 9: Guerinite, and 10: Portlandite. Red and black symbols (online version) represent minerals with and without As, respectively. (For interpretation of the references to colour in this figure legend, the reader is referred to the web version of this article.)

mobilities of As and Ca were positively correlated at pH 4 for Ci < 4 mmol Pi, their mobilities were not significantly correlated at pH 8 and 11. At pH 8 and 11, the mobility of As was independent of Pi concentration, whereas the mobility of Ca gradually decreased with increasing Pi. The SI values (Fig. S1) for several Ca–Pi–As(V) and Ca–Pi minerals were calculated at pH 4, 8 and 11, with the input solution compositions for these simulations having the As and Ca concentrations of the pH-specific non-Pi leachate, and the initial concentrations of Pi used in the sorption experiments. Therefore, the As and Ca concentrations at each pH were the same for all Pi concentrations used. Because Pi was not allowed to sorb on the soil during the PHREEQC simulations, the total Pi in solution was considered to be freely available to interact with aqueous As and Ca. The mineral SI values varied with pH, but for most minerals the SI values at each pH did not change significantly with increasing Pi. At pH 4 (Fig. S1a), PHREEQC simulations indicated that soil solutions were undersaturated with respect to all Ca–Pi–As(V) minerals. Similarly, these solutions were also undersaturated with respect to most of the Ca–Pi minerals (except amorphous Ca phosphate, hydroxyapatite and octocalcium phosphate). The calculated

mineral SI values at pH 8 and 11 were significantly different from SI values at pH 4. In particular, the simulated solutions at pH 8 were supersaturated with respect to several Ca–Pi–As(V) and Ca– Pi minerals (Fig. S1b). At pH 11, the solutions were approximately saturated with respect to most of the As-bearing minerals across the range of Pi concentrations used during the PHREEQC simulations. 3.3.3. Mobilization of As and Ca as functions of pH and Pi The mobilization of As and Ca as a function of pH in the presence of Pi was evaluated using the chemical data obtained from sorption experiments with initial Pi concentrations of 0.57 mmol and 3.38 mmol. Fig. 5 shows As and Ca concentrations when Pi reacted with BH soil between pH 3–12. The concentrations of Ca mobilized as a function of pH for 0.57 mmol and 3.38 mmol Pi were very similar; the highest mobility occurred at low pH. The mobility of Ca decreased rapidly with increasing pH, and became essentially immobile under circum-neutral to basic conditions. Arsenic also showed similar mobilization for both initial concentrations of Pi at pH > 6, but demonstrated significantly different mobilities for the two Pi concentrations at pH < 6. Under acidic conditions, As mobility was significantly higher for 3.38 mmol Pi. The lowest As mobility occurred between pH 6–9.5. At pH > 9.5, As mobility again increased until nearly 50% of the As in BH soil was leached at pH 12. PHREEQC modeling was also used to evaluate the potential precipitation of Ca–Pi–As(V) minerals as a function of pH from the nonPi soil leachate (Fig. 3) with the addition of 0.57 mmol and 3.38 mmol of Pi (Fig. S2). Both concentrations of Pi resulted in similar variations in the SI values for several minerals. The SI of each mineral increased with increasing pH and peaked between pH 7– 9; the SI values then decreased slightly at pH > 9. Between pH 6– 10, the simulated solutions were oversaturated with respect to most of the Ca–Pi–As(V) minerals, some Ca–Pi phases (e.g., octocalcium phosphate), and a few other minerals (e.g., hydroxylapatite) over the pH range from 3 to 12 (Fig. S2).

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Fig. 5. Percent mobilized soil As and Ca as a function of pH in the presence of 0.57 mmol or 3.38 mmol Pi. The percent mobilization was calculated with respect to the HNO3-digestible concentrations of these elements in BH soil (As = 3.71 mmol/ kg; Ca = 592 mmol/kg).

of externally supplied Ca, net transfer of Ca from solution to the soil occurred. At pH 8, nearly 45% and 60% of the externally supplied Ca was transferred to the soil, resulting in calculated Ca mobilities of 20% and 9% for 16.9 mmol and 5.63 mmol, respectively, of added Ca. With further increase in pH, a successively higher proportion of the externally supplied Ca was transferred to the solid phase. Almost 85% and 98% of the added Ca was transferred to the solid phase at pH 12 during Ca–Pi treatments using 16.9 mmol and 5.63 mmol Ca, respectively. The mobility of As also decreased with increasing pH (Fig. 7). At pH 8, about 3% and 10% of the HNO3-digestible As was released from BH soil after Pi treatments using 16.9 mmol and 5.63 mmol Ca, respectively. The Pi treatment with 5.63 mmol Ca resulted in a 4–6% mobilization of As at pH > 8. However, the Pi treatment

3.4. Phosphate and Ca treatment of soil 3.4.1. Sorption of Pi during Ca–Pi treatment Phosphate sorption on BH soil in the presence of externally supplied Ca was also evaluated for a Ci of 3.38 mmol Pi. For one set of experiments, 5.63 mmol Ca was supplied to yield a Ca:Pi ratio of 1.67:1 in solution; in a second set of experiments, excess Ca (16.9 mmol) was supplied, producing a Ca:Pi ratio of 5:1 in solution. Comparison of Figs. 6 and 2 shows that sorption of Pi on BH soil at pH < 5 was not significantly affected by external supply of Ca. At pH < 5, the sorption of Pi decreased with decreasing pH. Interestingly, the addition of 16.9 mmol Ca slightly decreased the sorption of Pi at pH < 5. However, the addition of external Ca significantly increased Pi sorption by the soil between pH 5–7, compared to Pi sorption in the same pH range without additional Ca. Moreover, the additional Ca resulted in a large increase in removal of Pi from solution at pH > 8. In the absence of external Ca (Fig. 2), increased, but incomplete, transfer of Pi to the solid phase was observed at pH > 7. However, almost 100% of the Pi was transferred from the soil solution to the solid phase under basic conditions when additional Ca was supplied (Fig. 6).

Fig. 7. Percent mobilized As and Ca as a function of pH in the presence of 3.39 mmol Pi and 5.63 or 16.9 mmol added Ca. The percent mobilization was calculated with respect to the HNO3-digestible concentrations of these elements in BH soil (As = 3.71 mmol/kg; Ca = 592 mmol/kg).

3.4.2. Mobilization of As and Ca during Ca–Pi treatments Fig. 7 shows the calculated mobilities of As and Ca as a function of pH when both Ca and Pi were supplied to the soil. The amount of Ca mobilized from BH soil was slightly higher across the pH range when less external Ca was supplied during treatment. In general, Ca mobility decreased with increasing solution pH (Fig. 7). No Ca was transferred from solution to the solid phase at pH < 5.8 and pH < 6.8 for 16.9 mmol and 5.63 mmol Ca, respectively. But at pH > 5.8 and pH > 6.8, respectively, for 16.9 mmol and 5.63 mmol

Fig. 6. Phosphate sorption on BH soil as a function of pH and initial Ca concentration. Initial Pi concentration was fixed at 3.38 mmol in all experiments, while the initial Ca concentration was either 5.63 mmol or 16.9 mmol.

Fig. 8. Mineral SI values as a function of pH calculated using PHREEQC: (a) 3.38 mmol Pi and 5.64 mmol Ca were added to each sample; (b) 3.38 mmol Pi and 16.9 mmol Ca were added to each sample. Mineral symbols – 1: Ca3(AsO4)2, 2: Ca3(AsO4)23.66H2O, 3: Ca3(AsO4)24.25H2O, 4: Ca4(OH)2(AsO4)24H2O, 5: Ca5(AsO4)3OH, 6: CaHAsO4H2O, 7: calcite, 8: Ferrarsite, 9: Guerinite, and 10: Portlandite. Red and black symbols (online version) represent minerals with and without As, respectively. Concentrations of aqueous As and Ca were taken from Fig. 3. (For interpretation of the references to colour in this figure legend, the reader is referred to the web version of this article.)

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with 16.9 mmol Ca consistently decreased the mobility of As with increasing pH. At pH 9, the mobility of As decreased to <1% of the HNO3-digestible As in BH soil. Figs. 8 and 9 show SI values for Ca–As(V) and Ca–Pi–As(V) minerals calculated by PHREEQC for the simulated solutions described above. Fig. 8 indicates that the addition of Ca alone was not adequate to trigger precipitation of Ca–As(V) minerals. The solutions were only supersaturated with respect to calcite at higher pH. The simulated solutions were undersaturated with respect to all Ca–As(V) minerals, even after the addition of 16.9 mmol Ca (Fig. 8b). However, the simulated solutions were oversaturated

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with respect to most of the Ca–Pi and Ca–Pi–As(V) minerals at pH > 5.5 (Fig. 9). 3.4.3. Characterization of precipitates Fig. 10 shows XRD patterns of precipitates separated after Ca–Pi treatment of BH soil at pH 8, and of material precipitated from the synthetic leachate. A reference XRD pattern for a carbonate apatite is also shown in Fig. 10. The XRD patterns for both precipitates closely match that of the carbonate apatite standard obtained from the ICDD database. 4. Discussion 4.1. Sorption of Pi on BH soil

Fig. 9. Mineral SI values as a function of pH calculated using PHREEQC: (a) 3.38 mmol Pi and 5.64 mmol Ca were added to each sample; (b) 3.38 mmol Pi and 16.9 mmol Ca were added to each sample. Mineral symbols – 11: Ca3(PO4)2, 12: Brushite, 13: Ca10(AsO4)0.42(PO4)5.58(OH)2, 14: Ca10 ðAsO4 Þ0:66 ðPO4 Þ5:34 ðOHÞ2 , 15: Ca10(AsO4)0.9(PO4)5.1(OH)2, 16: Ca10(AsO4)2.04(PO4)3.96(OH)2, 17: Ca10(AsO4)3.78 (PO4)2.22(OH)2, 18: Ca10(AsO4)5.64(PO4)0.36(OH)2, 19: hydroxylapatite, 20: CaHPO4, 21: CaHPO42H2O, 22: Ca4H(PO4)32.5H2O, 23: Ca3(PO4)2, and 24: whitlockite. Red and blue symbols (online version) represent minerals with and without As, respectively. Concentrations of aqueous As and Ca were taken from Fig. 3. (For interpretation of the references to colour in this figure legend, the reader is referred to the web version of this article.)

Fig. 10. Comparison of an XRD pattern of precipitates in Ca–Pi treated BH soil with an XRD pattern of material precipitated from synthetic Pi–As(V)–Ca leachate. The bar diagram at the bottom is a reference pattern for carbonate apatite obtained from the ICDD database.

Sorption of Pi is strongly dependent upon both pH and Ci; sorption is higher at acidic and basic conditions and lower at circumneutral pH, and greater fractions are sorbed at lower Ci. Sorption of oxyanion species, such as Pi, is generally high at low pH and decreases with increasing pH. Fig. 1 shows that the sorption of Pi on BH soil is highest at pH 4. The Freundlich isotherm model parameter, K, is much higher at pH 4 (51.5 L/kg) than at pH 8 (4.6 L/kg) or pH 11 (12.2 L/kg), which also reflects strong sorption of Pi under acidic conditions. Similarly, nearly 100% of the added Pi is sorbed by the soil between pH 3.5–5 for a Ci of 0.57 mmol (Fig. 2). In general, the sorption of Pi decreases with increasing pH from 4 to 7 (Fig. 2), which is attributed to adsorption controlled by pH-dependent surface charge. However, sorption of Pi is higher at pH 11 than at pH 8 until the equilibrium concentration (Ce) of Pi exceeds 6 mmol (Fig. 1). With increasing concentrations of Pi in the system, however, sorption of Pi at pH 8 also increases and approaches the level of sorption at pH 11 (Fig. 1). Fig. 2 also shows increasing sorption of Pi at pH > 7. Goldberg and Glaubig (1988a,b) reported similar observations for sorption of As(V) and Se(IV) on Ca-montmorillonite; however, greater oxyanion sorption with increasing alkalinity was absent when Ca was removed from the soil (Goldberg and Glaubig, 1988b). Therefore, the observed increase in sorption of Pi and other oxyanions at pH > 7 in the presence of Ca is interpreted to have been caused by precipitation. 4.2. pH controlled mobilization of As and Ca from BH soil Solution pH exerts a strong control over the mobility of As and Ca in BH soil. However, these elements have opposite mobility trends as a function of pH, except at low pH (3), where both elements display high mobility. The greater mobility of As at very low pH is caused by dissolution of As-rich minerals such as Al- and Feoxyhydroxides, as indicated by greater concentrations of Al and Fe in solution. In general, the mobilization of As increased with increasing pH from 4 to 12. The OH ligand is very effective in displacing oxyanion species such as As(V) from contaminated soil. Jackson and Miller (2000) found hydroxide to be the most effective anion for desorbing As, except As(III), from Fe oxides. Johnson and Barnard (1979) also reported that compounds rich in OH ligands are most effective in removing As from soil. Because Ca occurs as a cation, with low adsorption potential on metal oxyhydroxides at low pH (Rietra et al., 2001), it is highly mobile in acidic conditions. Moreover, acidic conditions increase the solubility of Ca-bearing minerals in soil, such as carbonates. However, at higher pH, Ca shows greater sorption (Rietra et al., 2001). On the other hand, adsorption of oxyanions on soil oxyhydroxide phases decreases with increasing pH. Arsenic in BH soil is mostly present as the oxidized form, As(V) (Yang and Donahoe, 2007); therefore, greater mobility of As with increasing pH is consistent with its aqueous geochemical behavior.

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The high mobility of Ca under acidic conditions is not accompanied by higher mobilization of As (except at pH 3), indicating that these elements are not associated with the same solid phase in BH soil. Using a 7-step sequential chemical extraction procedure, Qi and Donahoe (2008) reported that more than 60% of the HNO3-digestible As in BH soil is associated with amorphous Al– Fe-oxyhydroxides, and another 20% is associated with crystalline Al–Fe oxides. The remaining 20% was reported to be associated with the water soluble, exchangeable, carbonate, Mn-oxide, and organic soil fractions (Qi and Donahoe, 2008). Furthermore, the soil solutions are also undersaturated with respect to all Ca–As minerals (Fig. 3c). The relatively higher SI values calculated for all Ca–As minerals between pH 6–9 are due to the presence of both elements in soil leachate in this pH range. At pH < 6, the concentration of As is very low, and at pH > 8 the concentration of Ca is low. Therefore, the highly undersaturated state of the soil solution with respect to Ca–As minerals at low pH and high pH is controlled by low aqueous concentrations of As and Ca, respectively.

Similarly, As that leaches from BH soil does not precipitate, due to the apparent deficiency of Ca at pH > 8. Supplying adequate Ca in the treatment solution along with Pi would eventually lead to co-precipitation of As with Ca–Pi phases, and decrease its mobility. Similarly, the undersaturated state of the simulated solutions with respect to Ca–Pi–As(V) minerals at lower pH is caused by low concentrations of As, whereas at higher pH such undersaturated conditions are due to low concentrations of Ca in the non-Pi leachate. The greater mobility of As at low pH (except at pH  3) in the presence of Pi is caused by competitive desorption, rather than dissolution. However, the increased mobility of As at intermediate and higher pH is related to pH-induced desorption because As mobility at pH 8 and 11 is independent of Pi. Therefore, at circum-neutral to basic pH, the supply of additional Pi does not have a significant effect on the mobility of As. However, the addition of Pi increases the mobility of As under acidic soil conditions.

4.3. Phosphate induced mobilization of As and Ca from BH soil

4.5. Phosphate and Ca treatments

Although Pi solutions have been reported previously to be efficient agents for flushing As from soil (e.g., Peryea and Kammereck, 1997; Alam et al., 2001; Tokunaga and Hakuta, 2002), the effectiveness is dependent upon solution pH. The mobility of As at pH 4 increases with increasing initial concentration of Pi; however, As mobility is independent of Pi concentration in solutions at pH 8 and 11. Moreover, the mobility of As at pH 8 and 11 with varying amount Pi (Fig. 4a) is very similar to its mobility at these pH levels without Pi (Fig. 3). Therefore, the Pi-independent mobility of As at pH 8 and 11 can be attributed to pH-induced desorption of As associated with Al- and Fe-oxyhydroxides in the soil. The greater mobility of As at pH 4 is mostly related to Pi-induced competitive desorption. The Ca–Pi–As(V) and Ca–Pi mineral SI values calculated at pH 4, 8, and 11 by PHREEQC for simulated solutions having compositions obtained by adding the initial Pi concentrations to the non-Pi soil leachate, show different probabilities for mineral precipitation (Fig. S1). At pH 4 (Fig. S1a), the solutions are undersaturated with respect to all Ca–Pi–As(V) minerals. At pH 8, the simulated solutions are supersaturated with respect to several Ca–Pi–As(V) and Ca–Pi minerals (Fig. S1b), because the non-Pi leachate (Fig. 3) is slightly richer in Ca than the actual equilibrium solution (Fig. 4b). At pH 11, the solutions are approximately saturated with respect to most of the As-bearing minerals across the range of Pi concentrations. At pH 11, the controlling factor for the degree of saturation is presumably the amount of aqueous Ca, rather than the amount of As or Pi in solution. Hence, it is likely that the mobility of soil As can be decreased through precipitation of Ca–Pi–As(V) minerals, if adequate Ca is supplied along with Pi during soil treatment.

The sorption of Pi on BH soil at pH > 7 is changed when external Ca is also simultaneously added to the system. In particular, the additional Ca resulted in a large increase in removal of Pi from solution under alkaline conditions, and this nearly complete transfer of Pi from the treatment solution to soil is attributed primarily to precipitation. Precipitation of Ca–Pi phases under alkaline pH conditions is also supported by the observed negative mobility of Ca. A large fraction (45–60%) of the externally supplied Ca is transferred to the solid phase during soil treatment. The precipitation of Ca–Pi phases also helps decrease the mobility of soil As. In particular, the observed decrease in As mobility after Ca–Pi soil treatments (Fig. 7) is remarkable when compared to the high mobility of As during Pi-only soil treatments (Figs. 3a and b, 4a and 5). The decreased mobility of As, along with a large transfer of externally supplied Ca and Pi, indicates precipitation of As-bearing Ca–Pi mineral phases. PHREEQC simulations also showed that the solutions are oversaturated with respect to most of the Ca–Pi and Ca–Pi–As(V) minerals at pH > 5.5. Although the addition of 5.63 mmol and 16.9 mmol Ca with 3.38 mmol Pi resulted in very similar mineral SI values as a function of pH (Figs. 8 and 9), the BH soil treatment experiments using the lower Ca concentration (5.63 mmol) failed to completely immobilize the As. Therefore, the addition of excess Ca, along with Pi, is required for complete immobilization of As in contaminated soil. Similarly, speciation modeling indicated that the synthetic leachate (As: 0.48 mmol, Ca: 6.44 mmol, Pi: 3.38 mmol, pH: 8) equilibrated with atmospheric CO2 is oversaturated with respect to all of the Ca–Pi–As(V) minerals included in Table S2. However, the synthetic leachate is also undersaturated with respect to several Ca–As(V) minerals. Previous studies by Twidwell et al. (1994) and others (Wilson, 1998; Orser, 2001) have reported that removal of As from solution can be achieved by precipitation of apatite-like minerals at pH 8–10. Their studies also demonstrated that exposure of the experimental system to air (i.e., CO2) could potentially result in conversion of the precipitate into CaCO3, with subsequent release of As back into solution, if the Pi:As(V) ratio is < 7. However, if the Pi:As(V) ratio of Ca–Pi–As(V) minerals is P 7, the precipitates become stable under atmospheric conditions (Wilson, 1998; Orser, 2001). In the present study, the Ca–Pi–As(V) precipitate formed in BH soil after Ca–Pi treatment should have a Pi:As(V) ratio  7, as determined by mass-balance calculations. The precipitation of the Ca–Pi–As(V) solid phase, along with the low mobility of As, the negative mobility of Ca, and the nearly complete removal of Pi from solution at pH > 8, suggest that Ca–Pi treatment successfully immobilized As in BH soil.

4.4. Mobilization of As and Ca as functions of pH and Pi Mobilization of As and Ca as a function of pH in the presence of Pi is different, particularly at pH > 7. The increase in aqueous As concentrations without mobilization of Ca under basic conditions indicates that these two elements are not associated with a common phase in BH soil. Similarly, the lower mobility of Ca with higher As mobility, and the reversal in sorption of Pi at pH > 7 (Figs. 2 and 5), indicate precipitation of a Ca–Pi phase without incorporation of As. Although sorption of Pi by BH soil at pH > 7 is probably controlled by precipitation, only a portion of the total Pi added (up to 80% for 0.57 mmol and up to 40% for 3.38 mmol initial Pi concentration) (Fig. 2) is removed from the soil solution. The complete removal of Pi from solution by precipitation at high pH is prevented primarily by the deficiency of dissolved Ca.

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5. Conclusions In this study, the applicability of Pi treatment with and without added Ca for immobilization of As in contaminated soil was tested through various batch experiments. An As-contaminated soil sample collected from an industrial site was subjected to bench-scale Pi and Ca–Pi treatments. Batch experiments showed that the soil has a significant sorption capacity for Pi. Phosphate sorption as a function of the initial concentration of Pi and pH indicated that the least sorption of Pi, along with the lowest As mobility, occur at circumneutral pH conditions. Arsenic mobility increased while increasing the initial concentration of Pi at pH 4, but this trend was independent of Pi concentration at pH 8 and 11. Therefore, the mobilization of As from BH soil at pH 4 is related to Pi-induced competitive desorption, whereas mobilization of As at higher pH levels is attributed to pH-induced desorption. Treatment of BH soil with Pi and Ca resulted in decreased mobility of As. After Ca–Pi treatment, geochemical modeling indicated that soil solutions and synthetic leachate are oversaturated with respect to several Ca–Pi–As(V) phases. The externally supplied Ca and Pi are, therefore, most likely removed from solution by precipitation of solid phases during soil treatment. XRD analysis indicated the existence of a poorly crystalline carbonate-apatite phase in the Ca–Pi treated soil. A similar phase was also identified in precipitate from synthetic Ca–Pi–As(V) leachate. Most importantly, treatment of BH soil with 3.38 mmol Pi and 16.9 mmol Ca almost completely immobilized soil As from circum-neutral to high pH levels. The results of this study demonstrate that Ca–Pi treatment could be potentially applicable for immobilization of As in contaminated soils. In particular, this method would work best for As remediation in Ca-rich soils, which do not respond as well as low-Ca soils to other chemical fixation treatment methods. This treatment method could also be extended for remediation of Ascontaminated wastewater. The low solubilities of many of the Ca–Pi–As(V) minerals suggest that Ca–Pi treatment has promise as an effective, long-term method for in situ chemical fixation of As in contaminated soils and wastewaters. Acknowledgements The authors would like to thank Dr. Sidhartha Bhattacharyya, Elizabeth Y. Graham, and Ziming Yue for their help during laboratory experiments and chemical analyses. This research was partly funded by student research grants from the Geological Society of America, the Gulf Coast Association of Geological Societies, and the W.G. Hooks Fund (UA Department of Geological Sciences). Suggestions from three anonymous reviewers and Associate Editor Dr. Joyanto Routh were very helpful in revising the original manuscript. Appendix A. Supplementary material Supplementary data associated with this article can be found, in the online version, at http://dx.doi.org/10.1016/j.apgeochem.2012. 10.011. References Alam, M.G.M., Tokunaga, S., Maekawa, T., 2001. Extraction of arsenic in a synthetic arsenic-contaminated soil using phosphate. Chemosphere 43, 1035–1041. ASTM, 2003. ASTM D4567-03: Standard Test Method for Single-point Determination of Specific Surface Area of Catalysts and Catalyst Carriers Using Nitrogen Adsorption by Continuous Flow Method. ASTM International. ATSDR, 2000. Case Studies in Environmental Medicine – Arsenic Toxicity. ATSDR Publication No.: ATSDR-HE-CS-2002-0003 U.S. Department of Health and

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