Article pubs.acs.org/IECR
Remediation of Trichloroethylene by FeS-Coated Iron Nanoparticles in Simulated and Real Groundwater: Effects of Water Chemistry Eun-Ju Kim,† Kumarasamy Murugesan,† Jae-Hwan Kim,† Paul G. Tratnyek,‡ and Yoon-Seok Chang*,† †
School of Environmental Science and Engineering, Pohang University of Science and Technology (POSTECH), Pohang, 790-784, Republic of Korea ‡ Department of Environmental and Biomolecular Systems, Oregon Health & Science University, Beaverton, Oregon 97006, United States S Supporting Information *
ABSTRACT: The reactivity of FeS-coated iron nanoparticles (nFe/FeS) toward trichloroethylene (TCE) reduction was examined in both synthetic and real groundwater matrices to evaluate the potential performance of nFe/FeS in field treatment. The rate of TCE reduction increased with increasing pH, which is consistent with the pH effect reported previously for iron sulfide systems, but opposite that has been observed for (nonsulfidic) Fe0 systems. The rates of TCE reduction were unaffected by ionic strength over the range of 0.1−10 mM NaCl, increased with Ca2+ or Mg2+ concentrations, and inhibited by the presence of humic acid. The inhibitory effect of humic acid on the reactivity of nFe/FeS was largely alleviated when humic acid was combined with Ca2+/Mg2+, presumably due to decreased adsorption of humic acid onto nFe/FeS surface by the formation of humic acid−Ca2+/Mg2+ complexes.
1. INTRODUCTION In recent years, various methods have been established to enhance the reactivity of iron nanoparticles (nFe) for degradation of contaminants. The most common approach involves bimetallic combinations of nFe with catalytic, noble metals, such as Pd and Ni.1−3 However, the use of Pd or Ni increases the material cost and the catalytic effect of these noble metals is prone to poisoning under real environmental conditions, especially by reduced sulfur compounds.4 Recently, we reported an alternative enhancement strategy involving iron/iron sulfide core−shell nanoparticles (nFe/FeS), which exhibits better reactivity compared with nFe.5 Iron sulfide minerals commonly found in reduced, sulfidic groundwater and sediment have been shown to remove some contaminants due to reduction and/or adsorption.6−8 By coating nFe with FeS, we hope to gain some of the advantages of both types of materials. The reactivity of reducing materials, such as nFe and sulfide minerals, can be affected by their surface characteristics and the surrounding environmental conditions, including pH, ionic composition, and the presence of natural organic matter (NOM). NOM is ubiquitous in soils and natural waters and has a strong tendency to form surface complexes with iron and iron oxides through its carboxyl and hydroxyl functional groups. In most cases, it has been found that the presence of macromolecular NOM suppresses contaminant reduction by competition with organic contaminants for the reactive sites on the iron surface.9−12 Ionic strength and composition also play a role in the reactivity and longevity of iron-based reductants.13−16 Among them, the cations that characterize hard water (Ca2+ and Mg2+) have been reported to influence the contaminant removal alone and also in combination with carbonate species and NOM, presumably due to precipitation of insoluble reaction products.17,18 For example, when the Ca © 2013 American Chemical Society
ion is present together with NOM, it can promote NOM aggregation in solution via charge neutralization and compression of humic substances by the electrostatic binding of divalent cations. Thus, two or more of these solutes may interact with each other and then have combined influences that differ from the individual effects. Some prior studies have addressed the coupled impacts of solution components on contaminant reduction;19,20 however, to date, solution effects have mostly been examined individually. Although it has been shown that the reduction of several chlorinated solvents is significantly faster with nFe/FeS than conventional forms of nFe when performed in deionized water,5 further examination is required to evaluate the reactivity and stability of nFe/FeS under groundwater conditions. This paper aims to investigate the impacts of the major solution chemical variables on the reduction kinetics of trichloroethylene (TCE) by nFe/FeS. The synthetic water matrices were achieved by varying pH, monovalent (Na+) and divalent (Ca2+ and Mg2+) cation concentrations, and the presence of humic acid (representing NOM). The results are necessary for designing in situ remediation applications of nFe/FeS and understanding the reaction mechanisms in solution.
2. EXPERIMENTAL SECTION 2.1. Chemicals and Materials. Saturated stock solutions of TCE (≥99.5%, Sigma) were prepared in deionized (DI) water (Milli-Q). DI water was used throughout the entire experiment, and deoxygenated, deionized (DO/DI) water was prepared by purging with high-purity N2 for 1 h. The chemical Received: Revised: Accepted: Published: 9343
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apparatus (ELS-8000, Otsuka), and the EPMs were converted to zeta potential (ζ) based on the Smoluchowski’s equation. The hydrodynamic radius of the samples was also determined with ELS-8000 for dynamic light scattering (DLS) mode at a scattering angle of 90°. The precalculated amount of electrolyte solutions was added into the cuvette containing a diluted nFe/ FeS suspension (20 mg/L) and the measurements were started immediately. The early stage aggregation kinetics of nFe/FeS can be obtained from the initial rate of change of hydrodynamic size (Dh) with time (t). The aggregates/precipitates formed during the batch experiments were collected, and then dried in a vacuum oven for high-resolution transmission electron microscopy (HRTEM) (JEOL, Model JEM-2200FS) equipped for electron energy loss spectrometry (EELS), FTIR spectroscopy (Bomen MB104 spectrometer) and electrostatic force microscopy (EFM, Veeco Dimension 3100) analyses. FTIR spectra were measured with KBr pellets prepared by pressing mixtures of 1 mg sample and 100 mg KBr. For EFM measurements, the dried samples were also compressed into powder pellets using a hydraulic press, and the electrical leads were attached with silver paint onto the pellets.
stock solutions were prepared by dissolving chemicals (FeCl3·6H2O, NaBH4, Na2S2O4, NaCl, CaCl2·2H2O, and MgCl2·6H2O) into DO/DI water. All chemical reagents were of analytical grade and obtained from Sigma−Aldrich. Humic acid (HA) stock solution was prepared by dissolving Suwannee River humic acid powder (Standard II, International Humic Substances Society) in DI water, followed by filtration through a 0.45-μm cellulose acetate membrane (ADVANTEC) under vacuum. The concentration of HA was expressed as dissolved organic carbon (mg/L as DOC). A detailed description of the synthesis procedure of nFe/FeS has been reported elsewhere.5 Briefly, 150 mL of 0.017 M Na2S2O4 and 0.8 M NaBH4 was added dropwise (40−50 drops/min) to 50 mL of 0.5 M FeCl3. The resulting precipitate was retained with a strong magnet while washing three times with DO/DI water, and then dried in a vacuum oven at 50 °C for 1 d. 2.2. Batch Experiments and Analysis. Batch experiments were performed in 40-mL amber-colored glass vials capped with Teflon Mininert valves. These reactors were filled with 0.08 g of nFe/FeS and ∼40 mL of DI water/solutes or natural groundwater to obtain the 2 g/L suspension. Experiments were initiated by injection of 300 μL saturated aqueous TCE stock solution (for [TCE]0 ≈ 0.08 mM). The vials were then placed on a rolling mixer (15 rpm) at 26 ± 1 °C. Blank experiments without nFe/FeS were carried out in parallel, and there was negligible TCE loss. All experiments were done in duplicate. To examine the effect of pH (6.3−9) on TCE reduction rate, 0.1 M tris(hydroxymethyl)aminomethane (Tris)-buffer was used. Tris organic buffer was chosen because of the reported weak interaction with ferrous ions in solution. The pH of a buffer solution was adjusted to the desired value by adding small amounts of 0.1 M HCl or 0.1 M NaOH. In order to investigate the individual effects of ionic strength, Ca2+/Mg2+, and HA concentrations were varied over the ranges 1−10 mM, 0.1−2 mM, and 5−50 mg/L as DOC, respectively. Ionic strength was established with NaCl. The tested solute and HA concentrations were within the normal range of natural water to evaluate the potential of applying nFe/FeS in real aquifers.21,22 The combined effects of HA and hardness were evaluated at constant concentrations of Ca2+ or Mg2+ (0.8 mM) and HA (20 mg/L as DOC), respectively. Repetition experiments were also carried out in order to examine the long-term reactivity of the nanoparticles in real groundwater by exposing them to five sequential spiking of TCE (0.08 mM) at 3 h intervals. Herein, two different types of groundwater (GW1 and GW2) collected from rural areas in Korea were used and their physicochemical properties are given in Table S1 in the Supporting Information. The aqueous concentration of TCE was measured using headspace gas chromatograph equipped with an electron capture detector (GC-ECD) (HP Agilent, Model 6890), as described previously.5 The detection limit for TCE with this procedure was 0.05 μM. All data points indicate average values from duplicate samples, and error bars represent the standard deviation. Concentrations of HA were determined by a total organic carbon analyzer (Shimadzu TOC-V CSH ). Total dissolved Ca and Mg were quantified by an inductively coupled plasma-atomic emission spectrometry (ICP-AES, Thermo Jerrell Ash Corp.) 2.3. Particle Characterizations. The electrophoretic mobility (EPM) of nFe/FeS (20 mg/L) was measured as a function of pH using the electrophoretic light scattering
3. RESULTS AND DISCUSSION The detailed properties of nFe/FeS were reported elsewhere.5 TEM observations (Figure S1 in the Supporting Information) showed that the spherical particles in the range of 30-100 nm were aggregated forming chain-like structures. According to Xray photoelectron spectroscopy (XPS) analysis, the nFe/FeS contained ∼7.5 at. % sulfur (39.9 at. % Fe and 52.9 at. % O), which corresponded to sulfide (S2−) species. The specific surface area of nFe/FeS was calculated to be 39.7 m2 g−1 with a pore diameter of 4.3 nm. 3.1. Effect of pH. To investigate the effect of pH, the rate of TCE reduction by nFe/FeS was measured at pH 6.3, 7.0, 8.0, and 9.0 (Figure 1a). The kinetic data were fitted to the pseudofirst-order model based on previous papers,23,24 and the resulting observed rate constants (kobs) were tabulated in Table S2. The reduction rates clearly increased with increasing solution pH, which is consistent with previous observations in iron sulfide systems,7,25 but opposite the trend usually reported for Fe0 systems.26,27 The data gave a linear relationship between log kobs and pH over the pH range tested (Figure 1b), with a slope of 0.067 ± 0.003 (R2 = 0.99). This slopewhich represents the apparent reaction order with respect to [H+]is lower than the values previously reported for TCE vs nZVI26 or for TCE vs FeS.7 Taken together, these considerations suggest that the effect of pH on the FeS mediated reaction was attenuated by the opposing effects of pH on purely Fe0 and FeS phases. The pH-dependent behavior of nFe/FeS can be rationalized in terms of the equilibrium between specific FeS surface species in aqueous solution. The hydrated FeS surface contains iron hydroxyl (FeOH) and bisulfide (SH) functional groups, and these groups undergo protonation (FeOH + H+ ↔ FeOH2+) and deprotonation (SH ↔ S− + H+) reactions as the solution pH changes. The difference in reactivity between positively and negatively charged surface species is in part due to the greater driving force of deprotonated ligands in electron donation, which results in increased rates of TCE reduction at higher pH levels.7 Indeed, the cyclic voltammetry study performed on iron sulfides shows that peak currents 9344
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Figure 2. Zeta potentials (ζ) of nFe and nFe/FeS (20 mg/L) as a function of pH in 1 mM NaCl solution. Two different nFe/FeS particles synthesized with 0.017 and 0.028 M dithionite were used.
(Figure 3a) show that the particle sizes of nFe/FeS remained stable at low NaCl concentrations (1 and 10 mM), but increasing the concentration above 10 mM led to a marked increase in initial aggregation rate. In the presence of CaCl2, the aggregation occurred at lower concentrations than that
Figure 1. (a) TCE reduction by 2.0 g/L nFe/FeS at different pH values. The lines represent the fit to pseudo-first-order kinetics. (b) Logarithm of TCE reaction rate constant vs pH. The initial concentration of TCE was 0.08 mM.
increase with increasing pH, indicating that the rate of electron transfer reaction is faster at higher pH values.28 To further support our interpretation of the pH dependence of TCE reduction rates, in terms of FeS surface charge, the apparent zeta potential of nFe/FeS was determined as a function of pH. The isoelectric points (IEPs) for nFe and nFe/ FeS occurred at pH 7.3 and 4.2, respectively (see Figure 2). The low IEP value of nFe/FeS implies that a substantial part of the surface is covered by FeS, based on the reported values of IEP for iron sulfides (0.8−3.5)29 and Fe0 (7.0−8.0).30 It is notable that, despite varying the dithionite concentration used to make nFe/FeS (3.0 and 5.0 g/L), the IEP remains almost the same, indicating that the surface was effectively saturated with FeS so no further deposition of FeS was possible. Overall, these results show that nFe/FeS will have a large negative surface charge over the pH range of interest in most natural waters. 3.2. Aggregation of nFe/FeS in Monovalent and Divalent Electrolyte Solutions. Electrolyte composition and concentration can affect the aggregation of nanoparticles by altering the electrostatic energy barrier to attachment,31 which, in turn, can influence the reaction rate of organic pollutants in natural water environments.32 In order to measure the changes in hydrodynamic size of nFe/FeS over a range of monovalent (Na+) and divalent (Ca2+) electrolyte concentrations, we conducted the DLS experiments. The results
Figure 3. Aggregation profiles of nFe/FeS as functions of the (a) Na+ and (b) Ca2+ concentrations. Initial aggregation rates (in units of nm/ s) were as follows: Na+, 0.122 (20 mM), 0.242 (100 mM) and Ca2+, 0.129 (5 mM), 0.215 (10 mM). At NaCl 1 and 10 mM and CaCl2 0.1 and 1 mM, the aggregation rates were close to zero. 9345
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concentration of Cl− by adding NaCl did not increase the rate of TCE dechlorination (Figure 4). The lack of a stimulatory effect of Cl− on TCE reduction indicates that the kinetics of the TCE reduction reaction are not (entirely) limited by the kinetics of iron oxidation. This is not surprising, given the range of Cl− concentrations employed in this study, because previous work has shown that at least 25 mM Cl− is necessary to initiate breakdown of the passive film on micrometer-sized iron.36 However, follow-up experimentation that is beyond our current scope would be required to verify this hypothesis. The effect of HA on nFe reactivity is known to be either inhibiting or enhancing through various mechanisms, including competition with contaminants for the surface sites where reactions occur, formation of a protective and nonconductive layer, or mediated electron transfer.10,11 These effects might occur alone or in combination, depending on experimental conditions and the type of HA, so it is difficult to predict the net effect of HA on contaminant degradation. Using nFe/FeS, we found that increasing the initial HA concentration led to a decrease in the TCE reduction rate (Figure 4). The observed decrease in TCE reduction was attributed to the adsorption of HA (see Figure S3 in the Supporting Information), which competes with the contaminant for the reactive sites and/or forms a less-conductive film on the iron surface, thereby blocking TCE adsorption and reduction. Additional investigation of this effect using electrochemical methods showed that nFe/FeS gave the highest charge-transfer resistance in the presence of HA.34 The results suggest that the dominant effect of HA in this system is to form the organic coating that hinders both corrosion and contaminant reduction reactions. In order to examine the significance of the individual effects on nFe/FeS reactivity, we have correlated the kobs values and tested parameters. The correlation coefficients of pH, ionic strength, hardness cations, and HA were 0.99, 0.05, 0.99 (for Ca2+)/0.94 (for Mg2+), and 0.79, respectively. Except for ionic strength, all combinations exhibit a high degree of covariance and significance. 3.4. Combined Effects of Humic Acid and Water Hardness on TCE Reduction. The kinetics of TCE reduction were different when HA was present alone vs when HA was combined with Ca2+ or Mg2+ (Figure 4). The reaction rates followed the order: HA + Mg2+ > HA + Ca2+ > HA. The combination of HA and Ca2+ or Mg2+ gave TCE reduction rates that were ∼2−fold greater than solutions containing HA alone, implying that the presence of hardness cations overcomes the negative impact of HA in nFe/FeS system. This result differs from previous studies in which the inhibitory effect or no considerable difference in removal of targeted contaminants was observed in systems with a co-presence of NOM and Ca2+.19,20,37 The dissimilarities likely reflect the differences in the characters of NOM, solution pH, and pollutant types. It is well-established that adsorption and diffusion properties of NOM vary with its composition and surrounding conditions (i.e., pH and ionic strength).38 Therefore, the observed influence herein may be associated with changes in the sorption behavior of HA on nFe/FeS due to the complexation of HA with Ca2+ or Mg2+ in solution. The conceptual model showing the effect of hardness cation−HA complexation on TCE reduction is illustrated in Figure 5. The model shown in Figure 5 is further supported by evidence from EFM and FTIR analyses (Figure 6). The EFM results revealed lower brightness (see Figures 6a-I) for nFe/FeS
observed in systems with NaCl (Figure 3b), and a similar result was obtained with MgCl2 (data not shown). The substantially different effects of Na+ and Ca2+ on the aggregation of nFe/FeS is presumably due to the greater effectiveness of divalent cations in neutralizing the intraparticle electrostatic energy barrier via charge screening effects.33 However, on the basis of these findings, the range of electrolyte concentrations typically encountered in groundwater (1−10 mM for Na+ and 0.1−2 mM for Ca2+/Mg2+) is expected to have a negligible impact on the stability or the aggregation of nFe/FeS suspensions. 3.3. Individual Effects of Ionic Strength, Water Hardness, and Humic Acid on TCE Reduction. The kinetic results of TCE reduction obtained in the presence of various solutes (see Figure S2 in the Supporting Information) are summarized in Figure 4 and Table S2 in the Supporting
Figure 4. Rate constants of TCE reduction by 2.0 g/L nFe/FeS under various solutions. The initial concentration of TCE was 0.08 mM. The complete set of concentration versus time data was given in the Supporting Information (Figure S2).
Information. In this format, it is apparent that varying the ionic strength from 0.1 mM to 10 mM with NaCl had a negligible effect on the rate of TCE disappearance. In contrast, the presence of hardness cations (Ca2+ or Mg2+) significantly enhanced TCE reduction rates, and Mg2+ had a greater impact than Ca2+. These trends are consistent with differences in corrosion behavior of nFe/FeS, which we reported in a recent electrochemical study of this system.34 Under conditions similar to those used in the experiments reported here, it was found that nFe/FeS in NaCl solutions containing Ca2+ and Mg2+ showed higher corrosion currents than in DI water. The corrosion current for Mg2+ was slightly greater than that of Ca2+, suggesting that Mg2+ ions are more effective in promoting corrosion. In general, greater corrosion currents result from greater rates of iron oxidation at the surface, which implies greater reduction of TCE. On metallic iron, it is well-established that chloride promotes corrosion by favoring the formation of pits or crevices in the passive film,35 which may provide greater access of solution species to reactive surface sites, thereby increasing rates of contaminant dechlorination.9 From electrochemical measurement, it was shown that the system studied here (nFe/FeS) is even more sensitive to the corrosivity of chloride than nFe without FeS.34 However, in this study, increasing the 9346
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the blocking effect on access of TCE to reactive sites of nFe/ FeS. The difference in reactivity between the samples exposed to HA/Ca2+ and HA/Mg2+ may be attributed to the fact that Ca2+ has a higher binding affinity for HA than Mg2+.40 Thus, Ca2+ is more likely to form the complexes with HA in solution, so the relative contribution of Ca2+ to the surface adsorption and corrosion reaction will be smaller. In contrast, a comparatively large amount of Mg2+ ions can be adsorbed onto the nFe/FeS surface, which enhances the corrosion reaction. This interpretation is supported by EELS-based energy-filtered TEM observations (see Figure S4 in the Supporting Information). The map of Mg showed a stronger association with nFe/FeS particles than Ca. In addition, the fact that the samples exposed to HA/Mg2+ had a less compact morphology and low sulfur density indicates that the corrosion process was more severe under these conditions. Also, the final concentration of Mg2+ remaining in solutions was lower than Ca2+ (Figure S5 in the Supporting Information). 3.5. Implication for Performance under Field Conditions. As a preliminary evaluation of the stability and potential long-term performance of nFe/FeS in the field, TCE was repeatedly injected into batch reactors containing nFe/FeS and two different types of real groundwater (properties of GW1 and GW2 are given in Table S1 in the Supporting Information). The kinetics of TCE reduction were fit to pseudo-first-order model for each of five sequential respikes for both groundwaters (Figure 7). The values of kobs obtained in groundwater were smaller than in DI water containing Ca2+ or Mg2+ ions, presumably due to other dissolved organic or inorganic species (i.e., phosphate and carbonate) in these water samples. In each cycle, the kobs values for GW1 were slightly higher than those for GW2, which may be attributed to its greater hardness (see Table S1 in the Supporting Information). For both groundwaters, the kobs value decreased with each respike cycle, and the trends for the two water samples are roughly parallel (see Figure 7b). This could be due to the
Figure 5. Schematic illustration showing the possible effect of HA− hardness cation interaction on TCE reduction.
with HA only, compared with the sample that also contained Ca2+/Mg2+ (see Figures 6a-II and 6a-III). Assuming that the phase brightness in EFM is indicative to the electrical conductivity, the low surface conductivity of HA-amended nFe/FeS is in good accordance with its high charge transfer resistance observed by electrochemical impedance spectroscopy.34 In contrast, in the presence of hardness cations, nFe/FeS showed a relatively bright contrast (Figure 6a), indicating that HA adsorption was not significant. As seen in the IR spectra (Figure 6b), the typical peaks for HA were weak or not found in HA/Ca2+ and HA/Mg2+ samples. Indeed, in binary solutions of HA and Ca2+/Mg2+, apparently less HA adsorption was observed (Figure S3 in the Supporting Information). At higher pH, the dissolved HAs are able to form soluble complexes with Ca2+ or Mg2+ in solution that may adsorb poorly on the particle surface.39 Thus, formation of those soluble complexes results in less surface complexation of HA on nFe/FeS, thereby reducing
Figure 6. Surface analysis of the dried particles in solutions with HA (I), HA/Ca2+ (II), and HA/Mg2+ (III): (a) EFM images from 3 μm × 3 μm scanned areas and (b) FTIR spectra. The arrows in panel (b) refer to typical adsorption bands of HAs.41 Initial concentrations of HA and Ca2+/Mg2+ were 20 mg/L as DOC and 0.8 mM, respectively. 9347
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Figure 8. A logarithmic−logarithmic plot of surface-area-normalized rate constants (kSA, L m−2 h−1) versus mass-normalized rate constants (kM, L g−1 h−1) showing TCE reduction rates by nFe/FeS (squares). Previously published data for various Fe materials (without FeS) in DI water and artificial groundwater are shown as triangles.
4. CONCLUSIONS The present study provides valuable data on the reactivity of FeS-coated iron nanoparticles (nFe/FeS) for trichloroethylene (TCE) reduction under various water matrices, which will be necessary for designing operating conditions that achieve the most desirable performance. The stability and reactivity of nFe/ FeS was not significantly affected by the range of ionic strength that are often encountered in groundwater. On the other hand, hardness (Ca2+/Mg2+), and humic acid imposed contrasting effects on the particle reactivity. Although the presence of HA caused less than a 50% rate decrease, its concentrations used in the experiment are relatively high compared with those likely to be present in natural waters, implying that the inhibitory effects of HA would be small in real groundwater. Interestingly, the negative impact of HA was overcome when HA was present together with hardness cations. To the best of our knowledge, there have been no reports on this phenomenon and its positive effect on the reduction of organic contaminants. Based on these observations, it is revealed that nFe/FeS will be an attractive alternative over typical nFe for groundwater remediation. Further studies are needed to ensure the longterm performance of nFe/FeS as FeS can change to other iron sulfide phases with different surface reactivity over aging.
Figure 7. Repeated runs of TCE reduction in natural groundwater: (a) reduction of TCE by 2.0 g/L nFe/FeS and (b) rate constants as a function of sequential spike cycle. The initial concentration of TCE was 0.08 mM.
surface passivation of nFe/FeS by accumulation of reaction products or the presence of groundwater solutes. To provide further context for judging the significance of the effects reported here on the kinetics of TCE reduction by nFe/ FeSunder various hydrochemical conditions, including DI water/solutes and groundwaterwe have compared them to previously reported data for microscale and nanoscale elemental iron (Fe0) (see Figure 8).23,24 In this figure, the clusters along diagonal lines arise because the surface-areanormalized rate constants (kSA) is calculated from massnormalized rate constants (kM) by assuming one specific surface area value for each type of iron. The comparison suggests that nFe/FeS would have some kinetic advantages over various iron materials under field applications in several aspects: (i) most of groundwater solutes tested had a relatively small effect on the reduction rates of TCE by nFe/FeS compared with nFe, and (ii) the kM and kSA values of nFe/FeS in groundwater were generally about 1 order of magnitude higher than those of micro- and nanoscale Fe, indicating that the extent of groundwater inhibition in nFe/FeS system is somewhat less than that typically observed with Fe.
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ASSOCIATED CONTENT
S Supporting Information *
Tables of physicochemical properties of groundwater samples and summary of batch results. Figures of HRTEM images, DOC measurements, TEM/EELS images, and dissolved Ca/ Mg concentrations. This material is available free of charge via the Internet at http://pubs.acs.org.
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AUTHOR INFORMATION
Corresponding Author
*Tel.: +82-54-279-2281. Fax: +82-54-279-8299. E-mail:
[email protected]. Notes
The authors declare no competing financial interest. 9348
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ACKNOWLEDGMENTS This work was supported by the National Research Foundation of Korea (NRF) grant funded by the Korea government (MEST) (No. 2012-0008787) and “The GAIA Project” by Korea Ministry of Environment.
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