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Published online 27 October 2006
Published in J Environ Qual 35:2075-2083 (2006)
DOI: 10.2134/jeq2005.0467
© 2006 American Society of Agronomy, Crop Science Society of America, and Soil Science Society of America
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TECHNICAL REPORTS

Heavy Metals in the Environment

Bioaccessibility of Lead Sequestered to Corundum and Ferrihydrite in a Simulated Gastrointestinal System

Douglas G. Beaka, Nicholas T. Bastaa,*, Kirk G. Scheckelb and Samuel J. Trainac

a School of Environment and Natural Resources, The Ohio State University, Columbus, OH 43209
b National Risk Management Research Laboratory, USEPA, Cincinnati, OH 45224
c Sierra Nevada Research Institute, University of California-Merced, Merced, CA 95344

* Corresponding author (basta.4{at}osu.edu)

Received for publication December 19, 2005.

   ABSTRACT
TOP
 NOTES
 ABSTRACT
INTRODUCTION
 MATERIALS AND METHODS
 RESULTS AND DISCUSSION
 CONCLUSIONS
 REFERENCES
 
Lead (Pb) sorption onto oxide surfaces in soils may strongly influence the risk posed from incidental ingestion of Pb-contaminated soil. Lead was sorbed to model oxide minerals of corundum ({alpha}-Al2O3) and ferrihydrite (Fe5HO8·4H2O). The Pb-sorbed minerals were placed in a simulated gastrointestinal tract (in vitro) to simulate ingestion of Pb-contaminated soil. The changes in Pb speciation were determined using extended X-ray absorption fine structure (EXAFS) and X-ray absorption near edge spectroscopy (XANES). Both corundum (sorption maximum of 2.13 g kg–1) and ferrihydrite (sorption maximum of 38.6 g kg–1) have been shown to sorb Pb, with ferrihydrite having a very high affinity for Pb. The gastric bioaccessible Pb for corundum was >85% for corundum when the concentration of Pb was >200 mg kg–1. Bioaccessible Pb was not detectable at ≤200 mg kg–1. Bioaccessible Pb ranged from 53 to 88% for ferrihydrite. The bioaccessible Pb was below detection limits for the intestinal phase in the ferrihydrite system. Solid phase speciation identified both inner- (mononuclear bidentate) and outer-sphere species for Pb sorbed to corundum, while only an inner-sphere (mononuclear bidentate) complex was found for ferrihydrite. Although corundum and ferrihydrite can bind Pb, they fail to significantly reduce gastric bioaccessible Pb but do reduce intestinal bioaccessible Pb. Treatment of Pb-contaminated soil with corundum or ferrihydrite may reduce Pb solubility under field soil conditions of pH > 4. However, much of the sorbed Pb will become bioaccessible under gastric conditions (pH 1.5–2.5) if this soil is ingested. Caution should be used before using these materials to remediate a soil where soil ingestion is an important exposure pathway.

Abbreviations: DI, deionized • EDXA, energy dispersive x-ray analysis • GI, gastrointestinal • PZNPC, point of zero net proton charge • XRD, X-ray diffraction


   INTRODUCTION
TOP
 NOTES
 ABSTRACT
 INTRODUCTION
MATERIALS AND METHODS
 RESULTS AND DISCUSSION
 CONCLUSIONS
 REFERENCES
 
LEAD is a natural constituent in nearly all soils and geomedia (Hayes and Traina, 1998; Kabata-Pendias and Pendias, 2001). Soil concentrations of Pb are increasing from continual inputs from mining, agricultural sources, sewage sludges, fossil fuel combustion, metallurgical industries, electronic industries, chemical and other manufacturing, and waste disposal (Adriano, 1986; Basta et al., 2005). Because of increasing environmental Pb concentrations and the effects of Pb on children (Ryan et al., 2004) decreasing the Pb exposure to children is becoming increasingly important.

Incidental soil ingestion by children is an important exposure pathway in assessing public health risks associated with exposure to Pb-contaminated soils (Dudka and Miller, 1999; Ryan et al., 2004). The bioavailability of Pb in soils can be assessed by conducting dosing trials using animal models. The immature swine (Sus scrofa) has been used successfully as a model for gastrointestinal (GI) function of children (Ryan et al., 2004; Weis and LaVelle, 1991). There are several disadvantages to conducting in vivo animal trials such as length of time it takes to conduct the trial, expense of conducting the trials, and dosing problems. To overcome the difficulties and expenses associated with conducting in vivo trials to assess bioavailability of Pb in soils, research efforts have been directed toward the development of chemical in vitro methods. In vitro methods to simulate the gastrointestinal environment have been reviewed elsewhere (Oomen et al., 2002; Rodriguez et al., 1999; Ruby et al., 1999).

Before continuing, it is important to have an understanding of the terms bioavailability and bioaccessibility. For the purposes of this manuscript bioavailability will refer to solubilized Pb being transported across a biological membrane and entering systemic circulation. Bioaccessibility is the potentially available Pb solubilized from the test media.

Several in vitro assays have been developed to simulate human gastrointestinal absorption (Hack and Selenka, 1996; Minekus et al., 1995; Molly et al., 1996; Oomen et al., 2003; Oomen et al., 2002; Rodriguez et al., 1999; Ruby et al., 1993). In vitro assays that have been correlated with in vivo models include the in vitro gastrointestinal (IVG) (Rodriguez et al., 1999) and the physiologically based extraction test (PBET) (Ruby et al., 1993; Ruby et al., 1999). Both these methods simulate the GI tract conditions. Most of the difference between the two methods is in the gastric and intestinal fluid composition. Bioaccessible Pb by the IVG method is correlated with bioavailability to immature swine (Schroder et al., 2004).

Along with understanding the bioaccessibility of Pb in the environment, understanding the solid phase speciation of Pb in the environment is important. Lead speciation (Benson and Alberts, 1994; Erickson et al., 1992; Hayes and Traina, 1998; Hong et al., 1992) is key to understanding its fate and bioavailability. If the species present in the soil are easily solubilized then it is predicted that these species are bioavailable. Conversely if the species is insoluble then it is likely that the bioavailability will be low. This concept can be used in reverse to develop in situ remediation strategies to reduce the bioaccessibility of Pb in the environment.

It is well known that Pb interactions with hydrous oxides are important in moderating solubility and bioavailability. Hydrous oxides have been suggested as a possible amendment to remediate Pb in the soil (Berti and Cunningham, 1997; Chlopecka and Adriano, 1997; Lombi et al., 2002) and was used in combination with P to remediate Pb-contaminated soil in one study (Brown et al., 2004). These oxides are also natural sinks for Pb in the soil. Sorption reactions of Pb on metal oxides have been reviewed (Basta et al., 2005; Hayes and Traina, 1998). Aluminum oxides (Hsu, 1989) and Fe oxides (Schwertmann and Taylor, 1989) are the most abundant metallic oxides in the soil. The oxides of Fe and Al have a significant role in trace element retention, mobility, and bioavailability (Basta et al., 2005; Hayes and Traina, 1998). Because of this and their natural abundance, Al and Fe oxides make good candidates for Pb remediation in the soil.

The relationship between Pb surface speciation and bioavailability is unclear. To date there has been limited work relating the speciation of Pb to bioaccessibility. The research has been primarily examining the mineralogical and mineral solubility associated with in vitro models (Davis et al., 1992, 1996, 1999; Ruby et al., 1996). Past work primarily examined the gross mineralogical changes and associations using techniques such as X-ray diffraction (XRD) and energy dispersive X-ray analysis (EDXA) to determine the changes in the simulated gastrointestinal system. Lead in the environment is not always associated with a discrete mineral form. It can be sorbed to soil mineral surfaces and bound in co-precipitates/precipitates. The XRD and EDXA are not useful in determining speciation changes in highly heterogeneous environments. In soils that contain sorbed Pb, the solubility of discrete Pb minerals may not be as important as other soil chemical mechanisms (e.g., ligand exchange, surface properties, reductive dissolution, and proton assisted dissolution of the oxide mineral) in addressing the bioaccessibility of Pb.

Surface speciation of Pb sorbed to model Al and Fe oxides has been reported (Bargar et al., 1996, 1997a, 1997b, 2004; Ford et al., 1999; Scheinost et al., 2001; Strawn et al., 1998). For example, it has been shown that Pb will sorb to Al oxide surfaces as both inner-sphere and outer-sphere complexes (Bargar et al., 1996, 1997a; Strawn et al., 1998). Lead has been shown to form inner sphere complexes (Bargar et al., 1997b; Ford et al., 1999; Scheinost et al., 2001) and outer-sphere complexes (Bargar et al., 2004). For the inner-sphere Pb complexes, Pb is covalently bonded to the surface of the oxide, while the outer-sphere complex is held near the surface by electrostatic interactions These differences in bonding should relate to differences in bioaccessibility.

The risk posed from ingestion of Pb-contaminated soil is related to Pb bioaccessibility and may depend, in part, on sorption to oxide surfaces in soil. Knowledge of the relationship between Pb surface speciation and bioaccessibility of Pb in the GI tract will aid in understanding the risk posed by Pb-contaminated media. The objective of this study was to investigate the relationship between Pb bioaccessibility and surface speciation of Pb adsorbed to corundum, a model Al oxide and ferrihydrite, an Fe oxide mineral commonly found in soils.


   MATERIALS AND METHODS
TOP
 NOTES
 ABSTRACT
 INTRODUCTION
 MATERIALS AND METHODS
RESULTS AND DISCUSSION
 CONCLUSIONS
 REFERENCES
 
Ferrihydrite (Rhoton and Bigham, 2005) was obtained from a Memphis, TN, drinking water plant as a natural filtered by-product from the process of making drinking water. The ferrihydrite was washed three times with 18 M{Omega} deionized (DI) water to remove entrained solution. The solids recovered from the washing steps were then freeze dried and the dried solids were ground into a powder (<250 µm) using a mortar and pestle.

Corundum was obtained from Alcoa Industrial Chemical (Bauxite, AR). This corundum is the commercially available Tabular Alumina T-64-20 Micron product, and was formed by sintering alumina at 1800°C. The sintering process forms hexagonal tablet-shaped crystals with a particle size range of 40 to 200 µm. The purity was 99.7%.

Total elemental composition was determined using a CEM Microwave Digestion System (CEM MDS2100, Matthews, NC) using the CEM procedure for corundum (CEM, 1994). This method uses HF, HNO3, and HCl and pressure controlled (120 PSI) digestion for the complete dissolution of birnessite. The digests were analyzed for elemental content using a Varian Vista MPX (Varian, Palo Alto, CA) inductively coupled plasma–optical emission spectrometer (ICP–OES).

The XRD was used to confirm the identity of the ferrihydrite and corundum. The XRD data was collected from 10° to 70° 2{theta} using a Philips XRD diffractometer equipped with a graphite monochromatized Cu K{alpha} X-ray source (35kV, 20mA), 0.05° 2{theta} step size and a 4 s dwell time per step. Collected data was imported into the computer program, Jade (Materials Data, 1997) for data analysis and matched to the Powder Diffraction File database (International Centre for Diffraction Data, 1989).

Surface area was measured using the Brunauer–Emmett–Teller (BET) method (Carter et al., 1986) and a triple point N2 concentration adsorption isotherm. The surface area was determined from the slope of the regression through the three points. This was accomplished using Micromeritics FlowSorb II 2300 (Micrometrics, Norcross, GA) with the 2300 FC flow controller. The N2 concentrations used were 10, 20, and 30% N2 in He make up gas.

The point of zero charge (PZC) for each solid phase was estimated by measuring the point of zero net proton charge (PZNPC) (Zelazny et al., 1996). The method employed to measure PZNPC was a titrimetric method. The titration data was collected using a Mettler Toledo DL70ES titrator. The ferrihydrite was acidified (0.1 M HCl) to a pH below the estimated PZNPC (pH ~ 3.5) and a base (0.1 M NaOH) was added to an endpoint above the estimated PZNPC (pH ~ 9.5). For corundum, the suspension was acidified to a pH below the estimated PZNPC (pH ~ 5) and again a base was added to an endpoint above the estimated PZNPC (pH ~ 10.0). The ferrihydrite solid solution ratio was 5 g L–1 and the corundum solid solution ratio was 15 g L–1, the ionic strength in both solids of the sample was held constant with sodium nitrate (10–3, 10–2 and 10–1 M NaNO3) as the background electrolyte. Reproducible titration curves for corundum could only be obtained at solid solution ratios ≥15. Sample solutions and NaOH (0.1 M) titrant were made using CO2 free water. During the titration the solutions were kept CO2 free by bubbling N2 gas at a 0.5 mL min–1 flow rate. Triplicate analysis was performed for individual ionic strengths. The intersection of the titration curves indicated the PZNPC for the solids.

Concentrations of Al, Fe, and Pb were determined by ICP. Data for Al was collected using the 396.152 nm line for Al concentrations in the 0.01 to 10 mg L–1 range and the 394.401 nm line for Al concentrations in 10 to 200 mg L–1 range. The detection limit was 0.01 mg L–1. Iron was collected using the 238.204 nm line for Fe concentrations in the 0.01 to 10 mg L–1 range and the 261.382 nm line for Fe concentrations in 10 to 200 mg L–1 range. The detection limit was 0.01 mg L–1. Finally, Pb was collected using the 217.000 nm line which had a detection limit of 0.025 mg L–1 and was linear to 50 mg L–1. All standards and quality control samples were made in appropriate matrices. Quality control samples were run approximately every 30 samples and consisted of blanks and certified solutions (SPEX, Calibration Verification Standard, CL-ICV-1). Values for the certified solution were within ±10% of known values.

A Langmuir sorption isotherm was constructed for Pb sorption onto corundum and ferrihydrite (Sparks, 2003). This was accomplished by adding a 1 mL spike of a lead nitrate (Pb(NO3)2) standard at 10 concentrations (ranging from 0–10000 mg L–1 for corundum and 0–50000 mg L–1 for ferrihydrite) to a 20 mL suspension of corundum or ferrihydrite (6 g L–1 final solid concentration). Triplicate analysis was performed for each point on the isotherm. The temperature was held constant at 25°C, pH at ambient pH, and the suspension was placed on a reciprocating shaker (100 rpm) for 24 h. At the end of the shaking period, the suspension was centrifuged at 8000 rpm (8817 G) and the supernatant decanted. The supernatant was then analyzed for Pb using an ICP. The difference between the initial solution concentration and the final concentration was calculated. Based on this difference, the concentration of Pb sorbed to corundum and ferrihydrite could be calculated (Sparks, 2003; Sposito, 1989b).

Samples for bioaccessibility assays and speciation were prepared as follows. Corundum or ferrihydrite (3 g) was added to a 100 mL of solution containing the desired concentration of Pb and placed on a reciprocal shaker for 16 h. This suspension was then centrifuged at 8000 rpm (9820 G) for 15 min. The supernatant was poured off and filtered (0.45-µm membrane filter) for ICP analysis of Pb. The solids were resuspended with DI water to washout any entrained Pb solution. These solids were then centrifuged and the supernatant was poured off and filtered for ICP analysis of Pb. This wash step was repeated three times.

Lead bioaccessibility was determined by using the GI in vitro method of Rodriguez (Rodriguez et al., 1999) with modifications. Lead-treated corundum or Pb-treated ferrihydrite (3 g) was placed in a water jacketed reaction vessel containing 500 mL of degassed DI water, a variable speed mixer (Arrow Engineering), pH electrode, and gas dispersion tube. This suspension was equilibrated with Ar and thermally equilibrated to 37°C for 30 min. The mixing speed was 150 rpm and the Ar flow was 2 mL min–1. Upon thermal equilibrium, the gastric phase was initiated by lowering the pH of the suspension to 1.80 using 1 M HCl. The pH was maintained at 1.80 ± 0.05 for 120 min. The volumes of HCl additions were recorded for each addition. Subsamples (10 mL) were taken at 0, 5, 10, 15, 30, 45, 60, 90, and 120 min. These subsamples were immediately filtered using 0.45-µm membrane filters which mimic's the 0.5 µm particle size cutoff of human GI epithelial cells. The filtered solution was saved for ICP analysis. The intestinal phase was started on completion of the gastric phase (120 min) by raising the suspension pH to 7.0 using 1 M NaOH. The suspension pH was maintained at 7.00 ± 0.50 for 240 min. Additions of NaOH solutions were recorded. Sample aliquots (10 mL) were taken at 5, 10, 15, 30, 45, 60, 90, 120, 150, 180, and 240 min after initiation of the intestinal phase. The aliquots were filtered, preserved by the addition of HNO3 (pH < 2) and saved for ICP analysis. It is important to note that with constant mixing (both solution and solid was removed) the solid/solution ratio remained constant even though a significant volume of solution was removed.

Additionally, the total Pb content of the ferrihydrite or corundum was obtained using a CEM Microwave Digestion System and USEPA method 3051 (USEPA, 1994). This method is a microwave pressure (70 PSI) controlled digestion using HNO3. The digested solutions were filtered using 0.45-µm membrane filters and total Pb content was determined using ICP analysis.

Lead-sorbed ferrihydrite or Pb-sorbed corundum samples pre- and post in vitro assay were subjected to EXAFS and XANES to determine Pb surface speciation at the Advanced Photon Source (Argonne National Labs, Argonne, IL). The samples were loaded into the sample holder, sealed with kapton tape, and placed in the beam line. Lead L3 edge (13 035 eV) spectra were collected at the DuPont-Northwestern-Dow Collaborative Access Team (DND-CAT) sector 5, beam line 5BM-D using a Si 111 crystal monochromator and Canberra 13-element solid state detector (SSD) (Canberra, Meriden, CT). All spectra were collected at ambient temperatures. Raw spectra were averaged, normalized, background corrected, and converted to {kappa} space (EXAFS only) using Athena (Ravel and Newville, 2005). The {kappa} space data was then imported into WinXAS 3.1 (Ressler, 1998) for Fourier transformation into R space and shell-by-shell fitting of the radial structures. Theoretical paths for Pb, O, and Fe were calculated using Atoms (Ravel, 2001) and FEFF 6.0 (Mustre et al., 1990; Rehr et al., 1991, 1992). Data obtained from nonlinear least square included: coordination number (CN), bond distances (R), energy phase shift ({Delta}E0), and the Debye–Waller Factor ({sigma}2). For XANES spectra, the XANES spectra was imported into Win XAS 3.1 and derivatives or the spectra were taken. Additionally the EXAFS spectra were subjected to linear combination of fits (LCF) analysis to ascertain the possible phases present in the post in vitro samples. A more complete description of the fitting process can be found in Scheckel and Ryan (2004). A reference spectra library was created using the following model compounds: surface sorbed ferrihydrite and corundum, cerussite (PbCO3), pyromorphite (Pb5(PO4)3Cl, Pb acetate (Pb (C2H3O2)2·3H2O, simulate organically complexed Pb), Pb hydroxide (Pb(OH)2), massicot (orthorhombic PbO), litharge (PbO).


   RESULTS AND DISCUSSION
TOP
 NOTES
 ABSTRACT
 INTRODUCTION
 MATERIALS AND METHODS
 RESULTS AND DISCUSSION
CONCLUSIONS
 REFERENCES
 
The XRD analysis of ferrihydrite (Fig. 1A) showed two broad peaks at 35° 2{theta} and 62° 2{theta} which confirmed the presence of two-line ferrihydrite (Schwertmann and Cornell, 1991). However, the ferrihydrite XRD pattern also shows a sharp peak at approximately 26.5° 2{theta} which could indicate the presence of quartz in the sample. The quartz phase was confirmed by removing the Fe phase from the sample using three successive 6 M HCl washes and re-subjecting the remaining sample to XRD analysis (Fig. 1B). This not only confirmed the presence of the quartz phase, but also showed the presence of kaolinite in the ferrihydrite. The quartz and kaolinite content in the ferrihydrite is approximately 13% by mass. For corundum, the XRD analysis confirmed the presence of crystalline corundum (Fig. 2). The corundum XRD pattern did not indicate any other Al phases or impurities present.


Figure 1
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  		Fig. 1. X-ray diffraction pattern for ferrihydrite. A = unaltered ferrihydrite, B = iron phase removed.

 

Figure 2
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  		Fig. 2. X-ray diffraction analysis of corundum.

 
The elemental composition of the corundum and ferrihydrite are shown in Table 1. The elemental analysis confirmed the purity of the corundum, because the only major contaminant in the corundum was Na. However, the ferrihydrite used in this study has ambient levels of Pb and other trace elements (Table 1). There are significant amounts (g kg–1 levels) of Al, Ca, P, and Si contained in the ferrihydrite. This further demonstrates that the ferrihydrite contains other mineral phases (quartz and kaolinite) and strongly suggests the presence of other sorbed species such as phosphate and silica.


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  		Table 1. Elemental composition of corundum and ferrihydrite.

 
The surface area of the ferrihydrite was 229 m2 g–1. This is consistent with literature values reported for this solid (Goldberg and Johnston, 2001; Schwertmann and Cornell, 1991). The surface area of the corundum was 1.5 ± 0.06 m2 g–1, which is consistent with the manufacturer's analysis (1–2 m2 g–1).

The PZC of ferrihydrite was 4.5 ± 0.22 and is different from PZC of 7.0 to 9.0 reported in the literature for ferrihydrite (Dyer et al., 2003; Goldberg and Johnston, 2001; Schwertmann and Taylor, 1989; Zelazny et al., 1996). The presence of kaolinite (PZC 4.0–5.0) and quartz particles (PZC 2.0–3.0) (Sposito, 1989a) and/or sorbed P or Si anions (Goldberg and Johnston, 2001; Schwertmann and Taylor, 1989) gave a composite PZC that was less than typical for pure synthetic ferrihydrite. The PZC for corundum was 9.2 ± 0.05. This PZC value is consistent with what others have reported for corundum and other Al oxides (Davis and Kent, 1990; Hsu, 1989).

The sorption maximum (Pbmax) for Pb sorbed to corundum was found to be 2.13 g kg–1 (data not shown). Ferrihydrite however demonstrated a high sorption affinity for Pb which is reflected in the large Pbmax. The Pbmax for ferrihydrite was found to be 38.6 g kg–1 (data not shown). The Pb sorption site density for corundum and ferrihydrite was calculated from the sorption maximum and surface area. The Pb sorption site density for corundum was 4.13 sites nm–2, which is consistent with literature values. The site densities for Al- (hydr)oxides ranges from 2 to 12 sites nm–2, and the normal range is 1 to 7 sites nm–2 for natural minerals (Davis and Kent, 1990). The site density for ferrihydrite was 0.49 sites nm–2 which is lower than expected for ferrihydrite, 2.6 sites nm–2 (Davis and Kent, 1990). This lower Pb sorption site density may be due to the presence of sorbed elements (i.e., Ca, P, and Si) already bound to the ferrihydrite.

Bioaccessibility of Pb sorbed to corundum is shown in Table 2 and in Fig. 3A. Bioaccessible Pb was below the detection limit of <4 mg kg–1 when the concentration of Pb sorbed to corundum was <50 mg kg–1 (2% of Pbmax) for either the gastric or intestinal phases of the in vitro assay. However, when the sorbed Pb to corundum was ≥203 mg kg–1 (9.5% of Pbmax) the gastric phase bioaccessibility of Pb is >85% of the total sorbed Pb. The intestinal bioaccessible Pb was apparent, but highly variable and ranged from 3 to 24%. The bioaccessibility of Pb sorbed to ferrihydrite was slightly different and is shown in Table 2 and Fig. 4A. For all the ferrihydrite sorbed Pb concentrations, there was bioaccessible Pb in the gastric phase, but there was no measurable bioaccessible Pb in the intestinal phase. The bioaccessible Pb in the gastric phase ranged from 53 to 88%.


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  		Table 2. Bioaccessible Pb for Pb sorbed to corundum and ferrihydrite.

 

Figure 3
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  		Fig. 3. Bioaccessible (A) Pb and (B) Al for selected sorbed concentrations of Pb sorbed to corundum.

 

Figure 4
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  		Fig. 4. Bioaccessible (A) Pb and (B) Fe for selected sorbed concentrations of Pb sorbed to ferrihydrite.

 
A significantly greater mass of the solid was dissolved in the gastric phase for ferrihydrite (Fig. 4B). With the exception of the 41 700 mg Pb kg–1 sample (only 1% dissolved), 8 to 10% of ferrihydrite solid was dissolved in the gastric phase. Corundum dissolution was minimal with <0.2% of the corundum solid dissolved for all Pb concentrations (Fig. 3A).

Gastrointestinal absorption of Pb is a dynamic process involving dissolution of contaminant-containing mineral phases under strongly acidic conditions of the stomach and transfer of the contaminant from the small intestine into the systemic circulation. Biological GI digestive processes are quite complicated and difficult to simulate in vitro. In vitro GI methods based solely on measuring heavy metal contaminant solubility do not account for active and passive absorption processes and can only be accurate estimators of contaminant bioavailability if dissolution of the contaminant matrix is the rate-limiting step in this kinetic process (Basta et al., 2001; Ruby et al., 1999). Strong relationships from several studies between in vitro bioaccessible Pb or As and in vivo Pb or As suggest that dissolution of the contaminant matrix is the rate-limiting step for As and Pb in contaminated soils (Basta et al., 2001; Ruby et al., 1999). Bioaccessible Pb measured in the simulated gastric environment (e.g., gastric extraction step) is correlated with in vivo Pb (Ruby et al., 1996, 1999; Schroder et al., 2004).

Weaker relationships are found between bioaccessible Pb in the in vitro intestinal environment and in vivo Pb (Basta et al., 2001; Rodriguez et al., 1999; Schroder et al., 2004). The more acidic gastric phase of the in vitro GI methods extract more Pb and have a greater degree of precision than intestinal solutions where Pb is being removed from solution under nonequilibrium conditions. The gastric step of the in vitro GI methods is often used as a reproducible conservative estimate of in vitro bioaccessible Pb (Basta et al., 2001; Rodriguez et al., 1999; Schroder et al., 2004).

However, little information is available regarding the relationship between in vitro GI Pb and in vivo Pb (e.g., bioavailable Pb) for treated contaminated soils. It is unclear whether bioaccessible Pb measured under gastric or intestinal conditions is correlated with bioavailable Pb of treated soil measured by an appropriate in vivo model. Use of gastric bioaccessible Pb may be overly conservative whereas intestinal bioaccessible Pb may underestimate bioavailable Pb. The in vitro extraction (gastric or intestinal) selected to estimate bioavailable Pb in treated soil can profoundly affect the evaluation of the soil treatment method. Ferrihydrite greatly reduced bioaccessible Pb in the intestinal phase but had much less effect in the gastric phase (Fig. 4). Similarly, bioaccessible Pb was much less under intestinal conditions than gastric in the corundum system. Measurement of actual bioavailability of Pb in the Pb-spiked ferrihydrite or corundum systems requires a dosing study using an appropriate animal model and is beyond the scope of this work. Until in vivo data becomes available, the gastric bioaccessible Pb provides a conservative estimate of bioavailable Pb.

The EXAFS and fit parameters for Pb sorbed to corundum are shown in Fig. 5 and Table 3. The EXAFS results demonstrate that the surface speciation was the same for all Pb concentrations sorbed to corundum. There were no changes in Pb surface speciation before and after the simulated GI tract. The sorption of Pb on corundum occurs as inner-sphere and outer-sphere complex. The inner-sphere Pb-Al radius ranged from 3.34 to 3.36 Å, which correlates to a mononuclear bidentate complex (Bargar et al., 1997a; Strawn et al., 1998). The Pb–Al bond distance for the outer-sphere complex in the samples ranged from 5.76 to 5.78 Å, which is similar to what Barger has shown for corundum (Bargar et al., 1997a, 1996). The surface speciation of Pb sorbed to ferrihydrite was somewhat different than the surface speciation of Pb sorbed to corundum because there was no outer-sphere sorbed Pb as indicated by the EXAFS. Table 4 and Fig. 6 show the EXAFS fit data and EXAFS for Pb sorbed to ferrihydrite. The Fe–Pb bond radius ranged from 3.25 to 3.28 Å and is consistent with Fe–Pb bond radii observed by others (Bargar et al., 1997b, 2004; Ford et al., 1999; Scheinost et al., 2001). This bond radius is consistent with a mononuclear bidentate surface complex. As was the case with the Pb sorbed to corundum, there were no changes in speciation with respect to the Pb surface concentration. Also there were no changes in Pb surface speciation in the ferrihydrite samples taken before the in vitro assay and after the in vitro assay. Additionally, there was no change in Pb oxidation state during the simulated GI assay as shown in Fig.7. The Pb exists as Pb(II).


Figure 5
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  		Fig. 5. Extended X-ray absorption fine structure (EXAFS) spectra for Pb sorbed to corundum. Fits are shown as dotted lines.

 

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  		Table 3. Extended X-ray absorption fine structure (EXAFS) fit values for Pb sorbed to corundum before and after in vitro assay. CN = coordination number, R = radius, {sigma}2 = Debye–Waller Factor, {Delta}E0 = energy phase shift, IS = innersphere complex, and OS = outersphere complex.

 

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  		Table 4. Extended X-ray absorption fine structure (EXAFS) fit values for Pb sorbed to ferrihydrite before and after in vitro assay. CN = coordination number, R = radius, {sigma}2 = Debye–Waller Factor, and {Delta}E0 = energy phase shift.

 

Figure 6
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  		Fig. 6. Extended X-ray absorption fine structure (EXAFS) spectra for Pb sorbed to ferrihydrite. Fits are shown as dotted lines.

 

Figure 7
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  		Fig. 7. X-ray absorption near edge spectroscopy (XANES) comparison of Pb sorbed to corundum and ferrihydrite with model Pb compounds.

 
The LCF results are given in Table 5. In all cases the contribution of Pb (hydr)oxides was <10% and the majority of the Pb was surface sorbed Pb. The formation of pyromorphite in the 504 mg Pb kg–1 ferrhydrite sample is plausible since the ferrihydrite contained a significant P concentration. The dissolution of the P in the gastric phase and the subsequent raising of the pH could have initiated the formation of pyromorphite.


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  		Table 5. Results of linear combination of fitting for post in vitro samples.

 
A qualitative test was conducted to ascertain if simply raising the pH from gastric conditions (pH = 1.8) to intestinal conditions (pH = 7.0) would, in the absence of the solids, effect the solubility of Pb. Solutions containing Pb chloride (150 and 15 mg Pb L–1) and Pb nitrate (150 and 15 mg Pb L–1) were titrated from gastric pH using 0.1 M NaOH with constant mixing until precipitation occurred. Triplicate analysis was performed for each salt and concentration. For Pb chloride precipitation occurred at pH 5.33 ± 0.16 (125 mg Pb L–1) and 5.87 ± 0.24 (12.5 mg Pb L–1). A slightly higher pH was needed for the Pb nitrate solutions to precipitate, 5.92 ± 0.27 (125 mg Pb L–1) and 6.23 ± 0.09 (12.5 mg Pb L–1). This indicates that Pb precipitates are a possibility, however the EXAFS and LCF demonstrate that majority of Pb is re-adsorbed to the corundum or ferrihydrite and precipitation is a relatively minor.

The bioaccessibility of Pb sorbed to corundum is likely due to weakly bonded outer-sphere Pb complexes. Under gastric conditions, the surface of the corundum would have a significant positive charge due to the high PZC of 9.2 and low pH (pH = 1.8) of the gastric phase. The outer-sphere cationic Pb was Pb(II) which would tend to desorb from the positively charged corundum surface at pH 1.8. Under intestinal conditions, the corundum surface is less positive and would be less repulsive to sorption of Pb2+ compared to gastric conditions. There was some inner-sphere sorbed species of Pb(II) on the surface of the corundum. Low Pb bioaccessibility, found at low levels of Pb sorption, is consistent with inner-sphere sorption. Lead bioaccessibility increases with Pb loading because the weaker Pb outer-sphere complexes, found at higher Pb sorption, is easily desorbed at gastric pH of 1.8. Ferrihydrite, on the other hand, had inner-sphere bound Pb. It appears likely that the high dissolution of the ferrihydrite surface would explain the large bioaccessibility of the Pb. The mechanism is likely proton assisted dissolution of the ferrihydrite surface itself (Fig. 4B).


   CONCLUSIONS
TOP
 NOTES
 ABSTRACT
 INTRODUCTION
 MATERIALS AND METHODS
 RESULTS AND DISCUSSION
 CONCLUSIONS
REFERENCES
 
Both ferrihydrite and corundum could be used to sequester Pb in the environment. Ferrihydrite has a much higher affinity and sorption capacity for Pb than corundum and would be more useful in remediation of Pb contaminated media. Aluminum and Fe oxides have been used to immobilize Pb in contaminated soils. However, a large amount of Pb sorbed to these oxide minerals with be released under gastric conditions (pH 1.8). Ferrihydrite and corundum reduce bioaccessible Pb under much higher pH of intestinal conditions. Immobilization of Pb by Fe or Al oxides may not be a viable remediation technology to reduce site specific risk to human health when ingestion of contaminated soil is a critical exposure pathway when gastric bioaccessible Pb is used to estimate bioavailable Pb. However, intestinal bioaccessible Pb results suggest that treatment of Pb-contaminated soil with these oxides can significantly reduce risk associated with soil ingestion. Caution should be used before using these materials to remediate a soil where incidental ingestion is an important exposure pathway.


   ACKNOWLEDGMENTS
 
The authors thank Alcoa for supplying the T64-20 micron corundum research and Dr. Fred Rhoton for providing the ferrihydrite used in this research. Thanks to Jerry Bigham and Sandy Jones for the use and guidance on the XRD and surface area equipment in the Soil Characterization Lab in the School of Natural Resources at the Ohio State University. Use of the Advanced Photon Source was supported by the U. S. Department of Energy, Office of Science, Office of Basic Energy Sciences, under Contract no. W-31-109-Eng-38. Portions of this work were performed at the DuPont-Northwestern-Dow Collaborative Access Team (DND-CAT) Synchrotron Research Center located at Sector 5 of the Advanced Photon Source. DND-CAT is supported by the E.I. DuPont de Nemours & Co., The Dow Chemical Company, the U.S. National Science Foundation through Grant DMR-9304725 and the State of Illinois through the Department of Commerce and the Board of Higher Education Grant IBHE HECA NWU 96.


   NOTES
TOP
 NOTES
ABSTRACT
 INTRODUCTION
 MATERIALS AND METHODS
 RESULTS AND DISCUSSION
 CONCLUSIONS
 REFERENCES
 
Salaries and support provided by state and federal funds appropriated to the Ohio Agricultural Research and Development Center.


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





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