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Published in final edited form as: Environ Int. 2019 Feb 16;126:69–75. doi: 10.1016/j.envint.2019.02.022

Wheat straw biochar reduces environmental cadmium bioavailability

Liqiang Cui a, Matt R Noerpel b, Kirk G Scheckel b, James A Ippolito c,*
PMCID: PMC6487192  NIHMSID: NIHMS1525153  PMID: 30779992

Abstract

Cadmium contamination in soils and waters can lead to food chain accumulation and ultimately human health degradation; means for reducing bioavailable Cd are desperately required; biochars may play a role. Long-term lab incubation experiments aimed at explaining wheat straw-derived biochar (0, 5, and 15% by wt.) effects on Cd sorption, decreasing soil and solution Cd bioavailability, and identifying Cd forms present using both the European Community Bureau of Reference (BCR) wet chemical sequential extraction procedure and synchrotron-based X-ray absorption spectroscopy (XAS). Cadmium removal was 88 and 90% in the Cd-contaminated soil and Cd-containing solution, respectively, when using the 15% biochar application rate as compared to the control; significantly less Cd removal was observed with 5% biochar application rate. Based on the wet chemical sequential extraction procedure in conjunction with XAS, biochar application promoted the formation of (oxy)hydroxide, carbonate, and organically bound Cd phases. As a material, biochar may be promoted as a tool for reducing and sequestering bioavailable Cd from soils and solutions, potentially improving environmental and human health.

Keywords: BCR method, Biochar (BC), Cadmium, Contaminated paddy soil, Synchrotron, X-ray absorption spectroscopy

1. Introduction

Soil heavy metal pollution is a serious problem globally, leading to food chain accumulation and ultimately human health risk. In China specifically, soil heavy metal pollution can be linked to recent rapid industrialization, urbanization, and subsequently widespread solid, liquid, and gaseous metal emissions (Wang, 1998; Zhao et al., 2015). China’s Ministry of Environmental Protection (MEP, 2014) has estimated that almost 16% of all agricultural soils are contaminated with heavy metals. Of those heavy metals present, soil cadmium (Cd) concentrations exceed China’s environmental standards in 7% of all arable lands (98,000 km2) based on limits that protect agricultural production, heavy metal accumulation in the food chain, and consequently human health. When ingested, Cd may lead to kidney dysfunction, skeletal damage, and cancers (Satarug & Moore, 2004). According to the World Health Organization (2011), oral ingestion of as little as 0.125 mg Cd kg−1 body weight can cause health issues. Given the amount of arable land contaminated with Cd, the potential for human health issues is exacerbated in China.

A “cancer villages” map published online identified 247 locations in China with major health issues likely associated with man-induced pollution (China.org.cn, 2013). In China and other locations where soil heavy metal pollution is an issue, it is generally considered that human exposure to Cd is most often through the consumption of contaminated food (Zhao et al., 2013). In order to reduce the risk of heavy metal accumulation from soils, to plants, and ultimately humans, soil heavy metal bioavailability needs to be reduced. One such novel approach for reducing soil bioavailable metals is the use of biochar.

Biochar (BC), a carbon enriched material, is produced via pyrolysis of biomass at moderate temperature (~500°C) under limited oxygen concentrations (Lehmann et al., 2006). Because of its unique physical-chemical characteristics, BC is recognized as a multifunctional material for treating heavy metal polluted soils and water (Lehmann et al., 2011). Many studies (e.g., Devi & Saroha, 2014; Ippolito et al., 2017; Uchimiya et al., 2011) have reported that BC significantly sorbed heavy metals, leading to reduced heavy metal bioavailability and leaching. Multiple mechanisms of heavy metal sorption onto biochar have been proposed.

Ippolito et al. (2012) used synchrotron radiation to show that under slightly acidic or basic pH conditions, KOH steam-activated pecan shell biochar sequestered Cu as either bound to organic functional groups or precipitated as (hydr)oxides/carbonate phases, respectively. Lee et al. (2017) optimized pyrolysis conditions for Pb sorption onto palm oil sludge biochar. As with the Ippolito et al. (2012) findings, the authors showed that under acidic conditions the biochar sorbed Pb via cation exchange through various functional groups present. Peanut hull biochar was shown to be efficient in removing Cd (6.74 mg g−1) and Pb (63.09 mg g−1) from solution in the form of Cd carbonate (> 50%) and organo-Pb complexation (~40%) (Cui et al., 2014). Cui et al. (2016) utilized wheat straw biochar to reduce bioavailable soil Cd and Pb to more stable forms (e.g., residual forms), subsequently reducing Cd and Pb uptake by rice from contaminated paddy soil. Similarly, Bian et al. (2014) utilized a paddy soil contaminated with Cd and Pb from nearby smelter emissions, adding wheat straw biochar up to 40 Mg ha−1. Rice uptake of Cd and Pb were reduced by up to 69%, likely due to heavy metal interactions with surface functional groups, Fe oxyhydroxide phases, and Ca and P present in biochar. Ippolito et al. (2017) found that lodgepole pine or tamarisk biochar could reduce bioavailable Cd, Zn, Cu, and Pb concentrations by raising soil pH and subsequently forming carbonate and oxyhydroxide metal precipitates.

The above studies illustrate that biochars can reduce heavy metal bioavailability via a number of processes. However, the above studies also utilized either relatively short batch shaking periods (<24 hours) or relatively long field studies (3 to 5 years). We hypothesized that maximum biochar heavy metal sorption likely requires more than 24 hours to attain equilibrium, potentially on the order of days or weeks. Furthermore, we hypothesized that, given time, biochar-heavy metal interactions may lead from sorption via surface functional groups to precipitation of heavy metal mineral phases. Our objectives were to 1) utilize biochar to reduce bioavailable soil Cd and solution Cd, 2) investigate biochar-metal sorption and 3) identify form(s) present as affected by biochar application rate and time using both a sequential extraction procedure and X-ray absorption spectroscopy.

2. MATERIALS AND METHODS

2.1. Biochar and soil

Biochar was created from wheat straw pyrolyzed at ~450°C at the Sanli New Energy Company, Henan Province, China. After collection, BC was ground to pass through a 2-mm sieve.

The experimental soil was collected from a rice paddy field (31°24.434’N and 119°41.605’E) where atmospheric fallout and effluent discharges from a local smelter had contaminated the soil with Cd over several years. The paddy soil was air-dried and then passed through 2-mm sieve. Basic BC and soil properties are listed in Table 1 and were determined using methods described by Lu (2000). Briefly, pH was determined using a soil:water ratio of 1:2.5, cation exchange capacity (CEC) was detected via the ammonium acetate exchange method, total C using the potassium dichromate method, total N via Kjeldahl digestion and analyzed colorimetrically, total P via HClO4—H2SO4 digestion and detected using the molybdate blue method, total K via HClO4–HNO3–HF digestion and detected by atomic absorption spectrometry (FAAS; TAS-986, Persee, China), and total Ca, Mg and Cd using a HClO4–HNO3–HF digestion and analyzed using inductively coupled plasma–optical emission spectroscopy (ICP-OES, Perkin Elmer Optima 7300 DV, Shelton, USA).

Table 1.

Wheat straw biochar (BC) and paddy soil (0 to 15 cm depth) basic properties and total elemental analyses.

pH CEC† C N P K Ca Mg Cd
cmolc kg−1 g kg−1 mg kg−1
BC 10.4 21.7 467 5.90 14.4 11.5 10.4 6.49 0.03
Soil 6.1 18.0 20.7 3.19 0.82 11.4 1.46 5.38 22.6
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†

CEC = Cation Exchange Capacity.

2.2. Experimental design

2.2.1. Biochar Sorption of Cd from Solution

In triplicate, 0.00, 0.10, or 0.30g BC were weighed into 50 mL centrifuge tubes (equivalent to 0, 5, and 15% BC by wt.). Next, 40 mL of a Cd solution containing either 0, 5, or 15 mg Cd L−1, in 0.01 M KCl, was added to each tube; the solution pH was not buffered. The tubes were shaken at 120 rpm for 1h, 2h, 4h, 7d, 14d, 30d, 60d, 120d, and 240d. After each shaking period, the tubes were centrifuged (40 g) and solution pH was measured. The supernatant was filtered through a 0.45 μm membrane filter, two to three drops of concentrated HNO3 were added, and the solutions analyzed for Cd concentrations using ICP–OES. The biochar solid fraction was air-dried, crushed using a mortar and pestle, and then saved for X-ray absorption spectroscopy (XAS) analysis.

2.2.2. Biochar Sorption of Cd from Contaminated Soil

In quadruplicate, 2.00 g of contaminated soil and either 0.00, 0.10, or 0.30g BC (equivalent to 0, 5, and 15% BC by weight) were placed into 50 mL centrifuge tubes. Next, 40 mL of 0.01 M KCl (a proxy for metal bioavailability, as compared to Ippolito et al., 2017) was added to each tube; the solution pH was not buffered. The tubes were shaken at 120 rpm for 1h, 2h, 4h, 7d, 14d, 30d, 60d, 120d, and 240d. Following each shaking period, the tubes were centrifuged, the solution pH determined, and then the supernatant was filtered through a 0.45 μm membrane filter. Two to three drops of concentrated HNO3 were added, and the solutions analyzed for bioavailable Cd concentrations using ICP-OES. The biochar-soil solid fraction from 1h, 1d, 7d, 14d, and 120d were air-dried, crushed using a mortar and pestle, and saved for wet chemical-heavy metal sequential extraction and XAS analysis.

2.2.3. Biochar Effects on Soil Cd Fractionation (BCR Sequential Extraction)

The air-dried, crushed soil-BC composite samples from 1h, 1d, 14d, and 120d, were extracted using the four-step European Community Bureau of Reference (BCR) sequential extraction procedure according to Ippolito et al., (2017). Briefly, Step 1 (B1; soluble, carbonates, exchangeable fraction): 40 mL of 0.11 M acetic acid was added to 1 g of the soil-BC samples in a 50 mL centrifuge tube and shaken for 16 h at room temperature. The extract was then separated from the solid residue by centrifugation, decantation, and filtration through a 0.45 μm membrane filter. Step 2 (B2; iron and manganese oxyhydroxides fraction): 40 mL of a freshly prepared 0.1 M hydroxylamine hydrochloride was added to the residue from step 1 in the centrifuge tube and re-suspended by shaking for 16 h at room temperature. The extract was separated from the solid phase as described above. Step 3 (B3; organically bound and sulfides fraction): 10 mL of 30% H2O2 was added to the residue from step 2 in the centrifuge tube at room temperature for 1 h with occasional manual shaking, and then digested at 85 °C until the volume was reduced to about 3 mL. An additional 10 mL of 30% H2O2 was added and the digestion was continued at 85°C until the volume was reduced to 1 mL; the samples were allowed to cool. Finally, 40 mL of 1 M ammonium acetate was added and the tubes shaken for 16 h at room temperature. The extract was separated from the solid phase as described above. Step 4 (B4; residual fraction): the residue from step 3 was allowed to air dry and then it was digested in a 50 mL digestion tube using 1 mL deionized water to make a slurry, followed by aqua regia addition (7 mL concentrated HCl: 2 mL of concentrated HNO3). The mixtures were allowed to predigest overnight at room temperature and then digested at 105°C for 2 h the next day. Then, the mixtures were brought to a 50 mL final volume, filtered through a 0.45 μm membrane filter, and analyzed for Cd concentrations using ICP–OES.

2.2.4. Biochar-Cd Solution and Contaminated Soil X-Ray Absorption Spectroscopy (XAS) Analysis

Air-dried, crushed biochar from the 1h, 1d, 14d, and 120d Cd solution study, and air-dried crushed biochar amended soils from 1h, 1d, 7d, and 120d, were analyzed using XAS spectroscopy to elucidate solid-phase Cd speciation. The XAS experiments were conducted at the Materials Research Collaborative Access Team’s beamline 10-BM, Sector 10 located at the Advanced Photon Source, Argonne National Laboratory, Argonne, Illinois (Kropf et al., 2010). The electron storage ring operated at 7 GeV in top-up mode. A liquid N2-cooled, double crystal Si (111) monochromator was used to select incident photon energies, and a platinum-coated mirror was used for harmonic rejection. The samples were prepared as thin pellets with a hand-operated IR pellet press, and the samples were secured by Kapton tape in sample holders. Five XAS spectra were collected in fluorescence (Si drift detector) and transmission mode at room temperature from −200 to 800 eV relative to the absorption edge position of Cd. Data processing and linear combination fitting (LCF) of X-ray absorption near-edge structure (XANES) derivative spectra were done with the program Athena (Ravel & Newville, 2005).

2.2.5. Statistics

All data were expressed as means plus or minus one standard deviation. Differences between the treatments were examined using a Student’s t-test in a two-way analysis of variance (ANOVA), and considered significant when p < 0.05. All statistical analyses were carried out using SPSS (version 20.0, USA).

3. RESULTS AND DISCUSSION

3.1. Biochar Sorption of Cd from solution

As compared to the control, solution pH was significantly elevated over time (by ~1 to 4 pH units) with both 5 and 15% biochar applications in either 0 or 5 mg Cd L−1 (all p < 0.001) (Fig. 1A). When placed into the Cd solution containing 15 mg L−1, the 15% biochar application rate maintained about a 2 to 5 unit pH increase over the control over time (all p < 0.001); lesser increases were observed with the 5% biochar application rate. The biochar pH was 10.4 (Table 1), likely elevated due to the presence of oxides, hydroxides, and carbonate phases in the biochar (as shown by Ippolito et al., 2017). If biochar was removing Cd from solution, it may have occurred due to precipitation of Cd oxides, hydroxides, and carbonate mineral phases. Greater biochar application rates would have introduced greater oxides, hydroxides, and carbonate phases into solution, thus helping maintain an elevated pH over the control or lower biochar application rates. Concomitantly, increasing Cd solution concentrations would allow potentially allow for solid phase Cd precipitation reactions to occur, lowering solution pH as sorption sites and precipitation reactions occurred.

Fig. 1.

Fig. 1.

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Increasing wheat straw biochar (0, 5, and 15% by wt.) in solutions containing 0, 5, or 15 mg Cd L−1, on (A) solution pH and (B) residual solution Cd concentration after sorption onto biochar over shaking time.

Neither biochar application rate caused appreciable Cd loss from biochar itself when placed in solution containing 0 mg Cd L−1 (Fig. 1B), with Cd concentrations near or below detection. When placed in solution containing either 5 or 15 mg Cd L−1, there was always significantly less Cd in solution with the 15% versus 5% biochar application rate (p < 0.001). Over time, the 15% biochar application sorbed and removed practically all Cd from the 5 mg L−1 solution (>90% removal after 4 h), and removed almost 75% of the Cd from the 15 mg L−1 solution. As with the previous pH findings, as well as those of others (e.g., Ippolito et al., 2017; Xu et al., 2013; Kolodyńska et al., 2012), greater biochar application rates likely provided more sites for Cd sorption and precipitation reactions to occur.

Biochar dosage, contact time, and solution Cd concentration, all played a key role in Cd removal; others have found similar results. Bogusz et al. (2017) reported that the biogas residue biochar (600 °C) effectively removed Cd from solution, with sorption potentially affected by interfering anions (e.g., Cl−, NO3−). However, the authors suggested that Cd-(hydr)oxide precipitates blocked active sites on biochar surfaces, reducing interfering ion effects. Other inorganic ions (e.g., Ca2+, Mg2+, PO43-) from biochar were positively correlated to Cd removal from solution, promoting sorption or precipitation reactions (Clemente et al., 2017). Usman et al. (2016) also found that the palm waste biochar (700 °C) could remove almost 100% of Cd from solution, attributed to biochar functional groups (such as -OH, -CO32-) that promoted CdCO3 precipitation. Similarly, Štefelová et al., (2017) found that spruce sawdust wood biochar could sorb 13.4 mg Cd g−1, attributed to biochar surface functional groups (-OH) as pyrolysis temperature increased (400–700°C), promoting Cd precipitation reactions.

3.2. Biochar Sorption of Cd from Contaminated Soil

Increasing biochar application rate caused significant upward shifts in pH (p < 0.001) (Fig. 2A), likely due to the wheat straw biochar pH being 10.4 (Table 1). Additionally, when biochar is used as amendment, elements such as Ca, Mg, and K (Table 1) found on biochar sorption sites can be replaced with heavy metals such as Cd, causing a rise in soil solution pH (Huggins et al., 2016; Zambon et al., 2016). As suggested in the biochar-solution pH results above, biochar application likely supplied oxides, hydroxides, and carbonate phases to the soil and caused the pH to increase (Ippolito et al., 2017).

Fig. 2.

Fig. 2.

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Increasing wheat straw biochar (0, 5, and 15% by wt.) on Cd contaminated soil (A) pH and (B) bioavailable Cd concentration after sorption onto biochar over shaking time.

Biochar addition of 5 and 15% (by wt.) to the Cd contaminated soil reduced bioavailable Cd concentration by 53.4%−87.9% over the 240 d experiment as compared to the control soil (p < 0.001)(Fig. 2B). Supporting this result, Beesley et al. (2010) showed that an 8% biochar application rate (w/w) reduced the water-soluble Cd concentration in a multi-heavy metal contaminated soil in 56 days incubation pot experiment. A Chinese herb medicine biochar immobilized 81% Cd in contaminated soil following a 28 days incubation (Qiao et al., 2017). Lahori et al. (2017) reported that a 1% tobacco biochar significantly reduced plant-available Cd by 32% (Cd) over a control following ~ 10 weeks of incubation. Ippolito et al. (2017) mixed lodgepole pine biochar into four heavy metal contaminated soils, showing between 74 to 81% bioavailable Cd removal after only 2 h; the authors showed that Cd removal was due to the formation of Cd oxides, hydroxides, and carbonate phases as shown via the use of a BCR sequential extraction procedure.

3.3. Biochar Effects on Soil Cd Fractionation

Soil Cd fractions were operationally defined using the BCR sequential extraction procedure, with: step 1 (B1) associated with the soluble/carbonates/exchangeable fraction; step 2 (B2) associated with the iron and manganese oxyhydroxides fraction; step 3 (B3) associated with the organically bound and sulfides fraction; and step 4 (B4) associated with the residual fraction (Fig. 3). This approach is justifiable when the goal is to compare differences in a soil as affected by different treatments (Ippolito et al., 2009), or in the current study, by increasing biochar application rate.

Fig. 3.

Fig. 3.

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The effect of increasing biochar rate (0, 5, or 15% by wt.) over time (0.04 days [1h], 1, 14, and 120 d) on the ratio of (B1) soluble/exchangeable/carbonate, (B2) Fe/Mn oxyhydroxide, (B3) organically bound and sulfide fraction, and (B4) residual fractions of Cd in a Cd-contaminated soil.

In the current study, Cd increased in the soluble/carbonate/exchangeable fraction when 5 or 15% biochar was added as compared to the control (p < 0.001), while the 1 h, 1 d, and 14 d contained greater Cd in this phase as compared to 120 d (p < 0.001). There were no statistical treatment differences between Cd in steps 2 (iron/manganese oxyhydroxides fraction; p = 0.069) or 3 (organically bound/sulfide fraction; p = 0.365), while over time the iron/manganese oxyhydroxides fraction increased (p < 0.001). Cadmium content decreased in the residual fraction when 5 or 15% biochar was added as compared to the control (p < 0.001), while no differences over time were observed p = 0.965); this decrease may have contributed to the Cd increase in soluble/carbonate/exchangeable and the conversion of Cd bound in the iron/manganese oxyhydroxides fraction over time.

The increase in the soluble/carbonate/exchangeable fraction with biochar application might be attributed to Cd weakly sorbed onto the biochar surface or in the unstable biochar forms, as shown by Cui et al. (2014). Yet weakly sorbed or unstable forms would not likely not lead to a reduction in Cd bioavailability. Others (e.g., Egene et al., 2018; Cui et al., 2016) have shown water-soluble and exchangeable Cd reductions due to biochar application, but did not specify where Cd eventually resided.

However, others have identified Cd transformations from bioavailable to other soil forms. The reduction in Cd bioavailability has been associated with increasing soil pH and the formation of carbonate and oxyhydroxide precipitates when biochars were introduced into Cd-contaminated soil, as shown by Ippolito et al. (2017) using the BCR technique. Houben and Sonnet (2015) used an extraction procedure to show that miscanthus straw biochar shifted the Cd exchangeable pool to the carbonate pool primarily due to increasing pH. Moreno-Barriga et al. (2017) showed that biochar could reduce metal mobility due to increasing soil pH, and shift metals toward Fe/Mn oxyhydroxide phases. Shen et al. (2017) showed that biochar application could cause heavy metal shifts towards carbonate and carbonate-hydroxide phases. Thus, it appears that increasing and maintaining soil pH may help prevent dissolution and release of Cd back into the soil solution, which would re-promote Cd bioavailability.

3.4. Biochar-Cd X-Ray Absorption Spectroscopy (XAS) Analysis

Cadmium speciation in both the biochar-Cd solution sorption study and Cd in biochar amended soil investigation was determined through linear combination fitting (LCF) of the derivative XANES spectra. Four primary Cd species were identified, which include Cd associated with typical organic matter functional groups (OM/Biochar), cadmium sulfate (Cd Sulfate), cadmium bound to thiol (Cd Thiol), and Cd sorbed to Fe/Mn/Al (oxy)hydroxides or clay minerals (Mineral Bound).

Speciation results for the biochar-Cd solution sorption study were similar regardless of the biochar or Cd rate or as a function of treatment time. On average, OM/Biochar accounted for 66% (±2.16%), Cd Sulfate averaged 19% (±1.76%), and Cd Thiol measured 15% (±2.20%) (Table 2).

Table 2.

Speciation results of biochar sorption of Cd from solution.

Cd speciation distribution (%)
Biochar rate (%) Treatment time (d) OM/Biochar Cd sulfate Cd thiol Mineral Bound R-factor*
5 0.04 69 16 15 0.0021
15 66 18 16 0.0021
5 1 65 19 16 0.0032
15 62 20 18 0.0029
5 14 65 21 14 0.0019
15 66 20 14 0.0029
5 120 67 21 12 0.0078
15 68 20 12 0.0014
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*

A measure of mean square sum of the misfit at each data point.

Speciation results for biochar sorption of Cd from the contaminated soil varied relative to biochar rate and treatment time (Table 3, Figure S1). The control soil (0% biochar rate) showed consistent results as a function of treatment time, demonstrating that Cd was associated with 37±2.02% OM/Biochar, 17±3.15% Cd Sulfate, 20±2.18% Cd Thiol, and 26±1.68% Mineral Bound. In contrast, upon biochar amendment into soil there was a modest decrease in Mineral Bound Cd and Cd Thiol at each treatment time interval. Cd Sulfate showed a sharp increase at the 0.04-d treatment time relative to the 0% biochar value; however, Cd Sulfate in both the control and biochar treatments reduced in percentage over the 120-d reaction period to about 15%. Finally, the amount of Cd associated with OM/Biochar steadily increased as a function of treatment time, with slightly greater OM/Biochar values for the 15% biochar rate over the 5% rate.

Table 3.

Speciation results of biochar sorption of Cd from contaminated soil.

Cd speciation distribution (%)
Biochar rate (%) Treatment time (d) OM/Biochar Cd sulfate Cd thiol Mineral Bound R-factor*
0 0.04 37 21 18 24 0.0031
5 39 30 14 18 0.0022
15 38 29 19 15 0.0019
0 1 40 15 19 26 0.0011
5 46 20 13 22 0.0013
15 47 22 12 20 0.0017
0 14 37 16 20 27 0.0019
5 45 19 14 23 0.0023
15 50 17 11 23 0.0025
0 120 35 15 23 28 0.0015
5 48 15 13 25 0.0023
15 55 14 11 20 0.0037
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*

A measure of mean square sum of the misfit at each data point.

4. Conclusions

Some biochars are recognized for treating and removing heavy metals in contaminated waters and soils, reducing metal bioavailability and lessening metal effects on humans and the environment. We tested the ability of a wheat straw biochar, pyrolyzed at ~450 °C, to sequester Cd from solution and a contaminated soil. Based on wet chemical and XAS analysis, it appeared that functional groups present on wheat straw biochar helped to sorb, retain, and remove excess bioavailable Cd from solution and Cd contaminated soil. This particular form of Cd may be re-released back to the environment if biochar was to oxidize. Luckily, biochars pyrolyzed at modest temperatures (e.g., close to 500 °C) tend to be stable over the long-term (e.g., 100s of years), and thus this particular biochar may be beneficial for reducing solution and soil Cd bioavailability, and ultimately reducing bioavailable Cd exposure to humans and the environment.

Supplementary Material

1

Acknowledgements

This study was partially supported by National Natural Science Foundation of China under a grant number of 41501339, 21677119, 21277115, 41301551, 21407123, 41501353, Jiangsu Province Science Foundation for Youths under a grant number of BK20140468, sponsored by Qing Lan Project. MRCAT operations are supported by the Department of Energy and the MRCAT member institutions. This research used resources of the Advanced Photon Source, a U.S. Department of Energy (DOE) Office of Science User Facility operated for the DOE Office of Science by Argonne National Laboratory under Contract No. DE-AC02–06CH11357. Although EPA contributed to this article, the research presented was not performed by or funded by EPA and was not subject to EPA’s quality system requirements. Consequently, the views, interpretations, and conclusions expressed in this article are solely those of the authors and do not necessarily reflect or represent EPA’s views or policies.

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