Published in J. Environ. Qual. 33:1280-1287 (2004).
© ASA, CSSA, SSSA
677 S. Segoe Rd., Madison, WI 53711 USA
TECHNICAL REPORTS
Heavy Metals in the Environment
In Situ Speciation Studies of Copper–Humic Substances in a Contaminated Soil during Electrokinetic Remediation
S.-H. Liu and
H. Paul Wang*
Department of Environmental Engineering, National Cheng Kung University, Tainan City 70101, Taiwan
* Corresponding author (wanghp{at}mail.ncku.edu.tw).
Received for publication July 30, 2003.
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ABSTRACT
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Speciation of copper–humic substances (HS) in the electrokinetic remediation (EKR) of a contaminated soil was studied by in situ extended X-ray absorption fine structure (EXAFS) and X-ray absorption near edge structure (XANES) spectroscopies. The least-square fits of the XANES spectra suggested that the main Cu species in the contaminated soil were Cu–HS (50%), CuCO3 (28%), Cu2O (11%), and CuO (11%). The Cu–HS in the contaminated soil possessed equatorial and axial Cu–O bond distances of 1.94 and 2.17 Å with coordination numbers (CNs) of 3.6 and 1.4, respectively. In the EKR process, the axial Cu–O bond distance in the Cu–HS complexes was increased by 0.15 Å, which might be due to a ligand exchange of the Cu–HS with H2O molecules in the electrolyte. After 180 min of EKR, about 50% of the Cu–HS complexes (or 24% of total Cu) in the soil were dissolved and formed [Cu(H2O)6]2+ in the electrolyte, 71% (or 17% of total Cu in the soil) of which were migrated to the cathode under the electric field (5 V/cm). This work exemplifies the use of in situ EXAFS and XANES spectroscopies for speciation studies of Cu chelated with HS in the contaminated soil during EKR.
Abbreviations: CN, coordination number • EKR, electrokinetic remediation • EXAFS, extended X-ray absorption fine structure • FTIR, Fourier transform infrared • HS, humic substances • XANES, X-ray absorption near edge structure
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INTRODUCTION
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IN RECENT DECADES, soils contaminated by heavy metals have caused negative impacts to plant nutrition and food chains. In addition to their potential toxicity to human beings and other living organisms, heavy metals also affect ecological cycles under different redox and chemical environments (Giller et al., 1998). More than 160 heavy metal contaminated sites in Taiwan have been found (Sah and Chen, 1998). In the USA, about 5000 hazardous-waste-contaminated sites that may be leaking toxic chemicals into soil and then to the ground water have been reported (Khan and Alam, 1994).
Electrokinetic remediation, applying an electric field to the contaminated site, derives charged metal ions in the soil. The electromigration of ions and electroosmosis of water are major transport mechanisms during the ERK process. Electrokinetic remediation is one of the feasible technologies for in situ soil decontamination (Acar and Alshawabkeh, 1993; Lageman, 1993; Probstein and Hicks, 1993; Trombly, 1994). Like other in situ technologies such as bioremediation and soil flushing, EKR has advantages in avoiding the high cost and human health risks of excavation. In addition, EKR is very applicable in decontamination of heterogeneous and low-permeability soils. In lab-scale studies, EKR has been used successfully in removal of more than 90% of heavy metals (Cu, Cd, Co, Cr, As, Hg, Ni, Mn, Pb, Sb, and Zn) from clay, kaolinite, montmorillonite, and argillaceous sands (Acar et al., 1994; Eykholt and Daniel, 1994; Hamed et al., 1991; Hicks and Tondorf, 1994; Pamukcu and White, 1992; Yeung et al., 1996). Applications of EKR to contaminated soils in field trials also have been conducted recently (Lageman, 1993; Acar and Alshawabkeh, 1996; Ho et al., 1997). An average removal efficiency of Cu and Pb by EKR was about 74% (Lageman, 1993).
Although EKR has been proven to be very feasible in laboratory or bench-scale experiments and small-scale field tests, there is limited understanding of the complex transport phenomena and electrochemistry involved in the EKR. Many reactions in addition to electromigration and electroosmosis occur simultaneously when an electric field is applied to a wet soil. These reactions may include ion diffusion, ion exchange, mineral decomposition, precipitation of salts, hydrolysis, oxidation, reduction, physical and chemical sorption, and complexation with HS (Mitchell, 1993).
Humic substances including humic acids, fulvic acids, and humins are the most abundant groups of organic macromolecules in natural soils and water. Soil HS is a complex and heterogeneous mixture with large molecular weights of ubiquitous organic oligomers or polymers (Korshin et al., 1998; McIntyre et al., 1997; Sarr and Weber, 1980). Generally, soil HS contains several major functional groups such as carboxyl, phenolic, alcohol, and carbonyl. It is known that Cu–HS complexes play an important role in controlling the bioavailability and biogeochemical cycling of trace elements in the nature ecosystems. In addition, the mobility of Cu is strongly influenced by HS in soils. Studies of trace metal complexation with HS generally focus on measuring conditional stability constants and complexation capacities. However, few structural data for Cu–HS complexes have been reported (Lu and Johnson, 1997; Sarr and Weber, 1980). The abundance of different functional groups varies depending on the origin of HS and the methods of measurement. Many spectroscopic methods such as IR, fluorescence, nuclear magnetic resonance, and EPR have been used in revealing chemical structure of metal–HS complexes in contaminated soils (Fitts et al., 1999; Guthrie et al., 1999; Korshin et al., 1999; Paciolla et al., 1999).
Speciation data such as bond distance, coordination number (CN), and chemical identity of elements in the complex matrix can be determined by EXAFS spectroscopy. X-ray absorption near edge structure spectroscopy can also provide information of oxidation states of an excited atom, the coordination geometry, and the bonding of its local environment in environmental solids. By in situ EXAFS, we found that Cu oxides (in ZSM-5 or ZSM-48) were involved in the catalytic decomposition of NO (Huang and Wang, 1999; Huang et al., 2003) and oxidation of chlorophenols in supercritical water (Lin and Wang, 1999, 2000). Speciation of CuO during mineralization of CCl4 was also studied by in situ EXAFS spectroscopy (Chien et al., 2001). These in situ X-ray absorption data were very useful in revealing the speciation of Cu and the possible reaction pathways in the catalysis or mineralization processes.
Note that the main scientific issues concerning the speciation or chemical forms of contaminants ultimately depend on their molecular-scale structure. Basic understanding at this scale is essential in the management of environmental contaminants, which may also help the development of effective methods for remediation. Thus, the main objective of the present work was to investigate the speciation of Cu–HS in the EKR process by EXAFS and XANES. An in situ EXAFS cell was used to reveal the structural change of Cu during EKR. Additional chemical information of Cu–HS complexes in the contaminated soil was also investigated by Fourier transform infrared (FTIR) and X-ray photoelectron spectroscopy.
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MATERIALS AND METHODS
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Sample Preparation
The Cu-contaminated soil was sampled at a depth of 0 to 20 cm from a site that was near printed circuit board waste recycling plants in southern Taiwan. The soil samples were dried in air at room temperature for at least 3 d, ground with a mortar, and sieved with a 2-mm screen. Compositions of clay minerals in soil were identified by X-ray powder diffraction (XRD) (Model D/MAX III-V; Rigaku/MSC, The Woodlands, TX) spectroscopy with monochromatic CuK
radiation. The main minerals in the soil were kaolinite and mica (see Table 1). Concentrations of Al in the soil before and after EKR were determined by inductively coupled plasma emission spectrometry (ICP) (Model JY32/38; Jobin Yvon, Edison, NJ). Particle size analyses of the soil were conducted with the method described by Gee and Bauder (1986). The clay content in the soil was about 20% (see Table 1). The organic matter content, pH, and neutralizable acidity of the soil were determined using Page's method (Page, 1982). Humic acid (61%), fulvic acid (29%), and humin (10%) were found in the HS of the soil. The Cu–HS model compound was also prepared (Senesi et al., 1986) for XANES spectroscopic studies. About 50 mL of 0.01 M Cu(NO3)2 (Showa Chem, Tokyo, Japan) solution were well-mixed with 50 mL of a 10 g/L HS (Fluka, Buchs, Switzerland) solution at 300 K for 7 d. At pH = 2 (addition of a 0.1 M HCl solution), the Cu–HS was precipitated and separated.
Electrokinetic Remediation Apparatus
The EKR experiments were performed in a homemade in situ EXAFS cell (see Fig. 1) which consisted of two Ru electrodes, two electrode reservoirs, a power supply, a cation-selective membrane (Nafion 417; Aldrich, St. Louis, MO) (to prevent precipitation of Cu hydroxides on the electrode), and filters (Liu et al., 2001). About 60 g of the soil sample were filled uniformly in the cell and saturated with 0.01 M potassium nitrate as a conductive solution. A DC voltage of 100 V was constantly applied to the electrodes. After 180 min of EKR experiments, the soil in the cell was sliced into 10 portions of equal length immediately. Copper contents in each portion were determined by an acid digestion procedure (Page, 1982) and analyzed by flame atomic absorption (AA) (Model 5100; PerkinElmer, Wellesley, MA) spectroscopy.
X-Ray Photoelectron Spectroscopy
X-ray photoelectron spectra of the Cu-contaminated soil in the EKR process were measured on a Fison ESCA 210 spectrometer (VG Scientific/Thermo Electron, Waltham, MA) with a MgK X-ray (1253.6 eV) excitation source. The binding energy (Eb) of the Cu 2p3/2 line was calibrated with those of O 1s (530.8 and 532.5 eV) and C 1s (284.4 eV) lines. The standard deviation of the binding energy measurement was ±0.3 eV.
Fourier Transform Infrared Spectroscopy
The FTIR spectra of the soil–KBr pellets (wt. ratio of 1:25) were recorded on a Diglab FT–IR spectrometer (FTS-40) with fully computerized data storage and data handling capability. A 64-scan data accumulation was performed at a resolution of 4 cm–1.
Carbon-13 Solid-State Nuclear Magnetic Resonance
Carbon-13 solid-state nuclear magnetic resonance spectra of samples were also recorded at 9.4 T and 100.5 MHz on a Bruker (Billerica, MA) Avance DSX400 spectrometer. The soil samples were packed in 5-mm zirconia rotors and spun at a frequency of 5 kHz in the Bruker double-air-bearing probes. External tetramethylsilane (TMS) was used as a chemical shift reference for 13C. The recycle delay was 5 s with a pulse width of 10 µs (about 90° [flip angle]) for 13C.
In Situ X-Ray Absorption Spectroscopy
The in situ EXAFS cell could be moved precisely to obtain a similar optical path during X-ray absorption measurements. In the present study, the X-ray path was fixed onto a point 1.5 cm from the filter in the anode section (see Fig. 1). The Cu K-edge XANES and EXAFS spectra of the Cu-contaminated soil were collected (in the fluorescence mode) on the Wiggler beamline at the Taiwan Synchrotron Radiation Research Center (SRRC). The electron storage ring was operated at an energy of 1.5 GeV (current of 80–200 mA). A Si(111) double-crystal monochromator was used for selection of energy with an energy resolution of 1.9 x 10–4 (eV/eV). The absorption spectra were collected in ion chambers filled with helium gas. The beam energy was calibrated by the adsorption edge of Cu foil at an energy of 8979 eV. The standard deviation calculated from the averaged spectra was used to estimate the statistical noise and error associated with each structural parameter.
The EXAFS data were analyzed using the UWXAFS 3.0 and FEFF 8.0 programs (Ankudinov et al., 1998; Stern et al., 1995). The background of all the data was justified by the AUTOBK program (Ankudinov et al., 1998). The isolated EXAFS data was normalized to the edge jump and converted to the wavenumber scale. The Fourier transform was performed on k2–weighted EXAFS oscillations in the range of 3.5 to 10.5 Å–1. To reduce the number of fit variables, the many-body factor 
was fixed at 0.9. Generally, empirical fits of model compounds have an error of ±0.01 Å in radius and ±10% in CN for the first shell atoms, and ±0.02 Å and ±25% for the second shell atoms (Chisholm-Brause et al., 1990).
The absorption edge was determined at the half-height (precisely determined by the derivative) of the XANES spectra after pre-edge baseline subtraction and normalization to the maximum postedge intensity. The principal component analysis method (Fay et al., 1992) was used in the data treatment to optimize the quantitative extraction of relative concentrations of Cu species. The XANES spectra of model compounds such as the Cu–HS complex, aqueous Cu2+, Cu(NO)3, Cu(OH)2, CuSO4, Cu2O, CuCO3, and Cu foil were also measured on the Wiggler beamline. Semiquantitative analyses of the edge spectra were conducted by the least-square fitting of linear combinations of the model compound spectra to the spectrum of the soil sample. The height and area of the near edge band in a Cu spectrum were quantitatively proportional to the amount of Cu species.
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RESULTS AND DISCUSSION
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Because of the acid-induced dissolution of soil minerals (Wada and Umegaki, 2001), the major cation in the acid region may be Al ions, which may affect the speciation of Cu in the soil during EKR. Copper and Al concentrations and pH values in the soil between electrodes after EKR (180 min) were determined (see Fig. 2). Copper was accumulated near the cathode. The pH values on the anode and cathode were about 1.8 and 12.8, respectively. About 32 and 15% of Al and Cu, respectively, in the soil was dissolved in electrolyte near the anode during EKR. A similar dissolution trend for Al and Cu between electrodes was observed, suggesting an insignificant perturbation of Cu by the major cations (Al3+) during EKR.
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Fig. 2. Soil Cu and Al concentrations and pH values in the soil between electrodes after 180 min of electrokinetic remediation (EKR).
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Since the organic content in the contaminated soil was 8.1% (Table 1), it is also of great interest to determine the possible interactions between Cu and the organic matters. The main organic matter in the soil was HS. The FTIR spectra of the soil before and after the EKR treatments are shown in Fig. 3. The main features centered at 1616 and 1382 cm–1 can be assigned to HS carboxylate (Rhumic–COO–) asymmetric and symmetric stretchings, respectively. The absorption wavenumber differences (
) between the asymmetric and symmetric stretching vibrations of HS carboxylates (Rhumic–COO–) (which related to the mode of complexation [Nakamoto, 1978; Boyd et al., 1981]) are obtained from the IR data and shown in Table 2. For comparison, the absorption wavenumber difference of HS species was greater than that of the model compound Cu–HS by about 32 cm–1. Similar results were also observed for the complexation ( = 234 cm–1) of HS and Cu in the soil. After 180 min of the EKR treatments, the Cu–HS complexes in the contaminated soil were partially dissociated since a blue shift (1616
1658 and 15701612 cm–1) (see Fig. 3) of the COO– asymmetric stretchings as well as an increase of the absorption wavenumber difference were observed by FTIR spectroscopy.
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Fig. 3. Fourier transform infrared (FTIR) spectra of the Cu-contaminated soil (a) before and (b) after (180 min) the electrokinetic remediation (EKR).
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Table 2. Carboxylate (COO–) stretching wavenumbers of Cu–humic substances (HS) complexes in the electrokinetic remediation (EKR) process.
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In Fig. 4, the 13C solid-state nuclear magnetic resonance spectra of HS in the Cu-contaminated soil give additional chemical structure information. It is clear that aliphatic (0–50 ppm), oxygenated aliphatic (50–110 ppm), and aromatic (100–160 ppm) carbons of the HS were perturbed insignificantly by EKR. However, the feature centered at 173 ppm that was attributed to carboxylic carbons (160–220 ppm) was slightly reduced in intensity in the EKR process, suggesting the possible dissociation of the Cu–HS complexes (as observed by FTIR).
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Fig. 4. Carbon-13 solid-state nuclear magnetic resonance spectra of the Cu-contaminated soil (a) before and (b) after (180 min) the electrokinetic remediation (EKR).
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Surface compositions of Cu in the contaminated soil with the EKR treatments were also studied by X-ray photoelectron spectroscopy. The Cu 2P3/2 spectrum of the contaminated soil (Fig. 5) suggests that Cu on the surfaces of contaminated soil contains at least two oxidation states [Cu(II) and Cu(I)]. Note that a small amount of the Cu(I) species originally in the contaminated soil was oxidized to Cu(II) in the EKR process. An intense broad feature at 943 eV is attributed to the so-called shake-up satellite and unmistakable feature of open-shell d9 species [Cu(II)]. The d shells for Cu(0) and Cu(I) species are completely filled with 10 electrons and the satellite feature is almost completely vanished.
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Fig. 5. X-ray photoelectron spectroscopy spectra of the Cu-contaminated soil (a) before and (b) after (180 min) the electrokinetic remediation (EKR).
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To further understand the structural changes of Cu–HS during EKR, oxidation state, bond distance, and CN of Cu were studied by in situ XANES and EXAFS. The in situ XANES spectroscopy, being generally very sensitive to changes of the ligand environment, can provide information about how electron transitions and multiple scattering are influenced in the EKR process. The in situ Cu K-edge XANES spectra and their first derivatives of the Cu-contaminated soil during EKR are shown in Fig. 6. The pre-edge XANES spectra of Cu in the contaminated soil exhibit a very weak 1s-to-3d transition that is forbidden by the selection rule in the case of perfect octahedral symmetry (Iwamoto et al., 1998; Liu et al., 2001; Yamashita et al., 1996). The pre-edge band at 8982 eV may be due to the dipole-allowed 1s-to-4p transition of Cu(I). At the early stage of EKR (see Fig. 6), the content of Cu(I) was, to some extent, decreased, which might be due to oxidation of Cu involved in the EKR process. The first derivatives of the XANES spectra revealed two shoulders at 8986 eV () and 8992 eV (ß). Generally, the ß feature represents the main absorption transition (1s-to-4p) while the feature is influenced by the degree of bond covalency and the degree of the local structure disorder (Palladino et al., 1993). The observation of the and ß features suggested the existence of Cu(II). Prolonging the EKR time to 180 min, a slight perturbation of the and ß features of Cu in the soil was observed. Frenkel et al. (2000) also observed similar results for the Cu–HS species.
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Fig. 6. In situ X-ray absorption near edge structure (XANES) spectra and their first derivative of the Cu-contaminated soil treated by electrokinetic remediation (EKR) for (a) 0, (b) 60, (c) 120, and (d) 180 min. Solid lines and circles represent experimental data and the least-square fits, respectively. Dotted lines denote fractional contributions of the five main components (Cu–humic substances [HS], CuCO3, CuO, Cu2O, and aqueous Cu2+) making up the fitted spectra.
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The XANES spectra were also expressed mathematically in a LC XANES fit vectors, using the absorption data within the energy range of 8970 to 9020 eV. The XANES spectra of model compounds such as Cu–HS complex, aqueous Cu2+, Cu(NO)3, Cu(OH)2, CuSO4, Cu2O, CuCO3, and Cu foil were also measured on the Wiggler beamline. It was found that Cu–HS, CuCO3, CuO, and Cu2O were the main Cu species in the Cu-contaminated soil. Relative contents of the Cu species in the EKR with more than 90% reliability (in the data fitting process) are shown in Fig. 6. It is clear that perturbation and dissociation of Cu–HS complexes affected by electric field might lead to a decrease of about 50% of the Cu–HS species in the soil. Oxidation of Cu(I) was also observed during EKR. However, CuCO3 was perturbed insignificantly after 180 min of the EKR treatments. It should be noted that CuCl, CuCl2, Cu(NO)3, and Cu(OH)2 were not identified in the in situ XANES spectra, suggesting an insignificant perturbation of anions during EKR. The time dependence for dissolution and migration of Cu in the soil during EKR is shown in Fig. 7. The dissolution and electromigration rates of Cu(II) in the soil were 14.5 and 8.2 µmol/(h g) soil, respectively, during EKR. Migration of Cu to the cathode might be retarded because of the readsorption of Cu onto surfaces of the soil.
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Fig. 7. Time dependence for (a) dissolution and (b) electromigration of Cu in the soil during electrokinetic remediation (EKR).
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The in situ EXAFS spectra were also recorded and analyzed in the k range of 3.5 to 11.5 Å–1. Due to the axial distortion of the CuO6 octahedron caused by the Jahn–Teller effect (Palladino et al., 1993), the distances between the Cu and the equatorial and axial oxygens were not identical. The major feature at 1.56 to 1.60 Å (uncorrected for phase shift) is attributed to the first-shell Cu bonded to oxygen atoms (equatorial Cu–O bonding) (see Fig. 8). The feature centered between 2.18 and 2.35 Å represented scattering from additional oxygen atoms (axial Cu–O bonding) in the first coordination shells. Different axial and equatorial Cu–O bond distances and values of Debye–Waller factor (
2) for the EKR of Cu-contaminated soil are also shown in Table 3, which also provides important information of the inner shells in complexation of Cu with HS. The Cu–HS in the contaminated soil had equatorial and axial Cu–O bond distances of 1.94 and 2.17 Å with CNs of 3.6 and 1.4, respectively. Interestingly, the equatorial and axial Cu–O bond distances of the Cu–HS in the soil were increased in the EKR process. Generally, the d orbitals contain double (eg) and triple (t2 g) degeneracies for transition metals in the octahedral field. For d9 metals such as Cu(II), only three electrons occupy the higher-energy eg level. While the degeneracy of the eg set of orbitals (dz2 and dx2–y2) is removed, one of the two orbitals is stabilized by lengthening either two axial or four equatorial metal ligand bonds (Huheey, 1978). Weak-field equatorial ligands (such as COOH) of Cu can be replaced by stronger ligands (H2O). An increase in the axial Cu–O bond distances (from 2.17 to 2.32 Å) in the soil was found in the EKR process, which might be due to the ligand exchanges in the EKR process. Carboxylic ligands of HS may be replaced by water molecules during the EKR. After 180 min of EKR, about 50% of the Cu–HS complexes were dissolved in the electrolyte.
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Fig. 8. Normalized Cu extended X-ray absorption fine structure (EXAFS) k2–weighted chi (left panel) and corresponding Fourier transforms (right panel) of the Cu-contaminated soil treated by electrokinetic remediation (EKR) for (a) 0, (b) 60, (c) 120, and (d) 180 min. Dotted and solid lines denote the fitting and experimental data, respectively.
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Table 3. Speciation changes of Cu–humic substances (HS) in the contaminated soil in the electrokinetic remediation (EKR) process.
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The possible reaction path for Cu–HS complexes involved in the EKR process is shown in Fig. 9. In the EKR process, transformation of the (Rhumic–COO)4–Cu complexes into four Rhumic–COOH and aqueous Cu(II) (possibly [Cu(H2O)6]2+) was observed by combined FTIR and in situ XANES and EXAFS spectroscopies. The increases of the axial Cu–O bond distance (determined by in situ EXAFS) in the EKR process may be due to the ligand exchange reaction in which carboxylic ligands of HS in the equatorial plane of the Cu(II) species may be replaced by the strong-field water molecules.
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Fig. 9. The possible reaction path for Cu–humic substances (HS) complexes involved in the electrokinetic remediation (EKR) process.
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CONCLUSIONS
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By the least-square fits of the XANES spectra, fractions of main Cu species such as Cu–HS (50%), CuCO3 (28%), Cu2O (11%), and CuO (11%) were found in the contaminated soil. The in situ EXAFS data indicated that the averaged equatorial and axial Cu–O bond distances in the Cu-contaminated soil were 1.94 and 2.17 Å with CNs of 3.6 and 1.4, respectively. However, an increase in the axial Cu–O bond distances (2.172.32 Å) was found in the EKR process. The observations may be due to the possibility that weak-field carboxylic acid groups in the equatorial plane of the Cu(II) species were replaced with the strong-field water molecules. After 180 min of EKR, about 50% of the Cu–HS complexes (or 24% of total Cu) in the soil was dissolved into the aqueous phase, 71% (or 17% of total Cu in the soil) of which was migrated to the cathode under the electric field (5 V/cm).
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ACKNOWLEDGMENTS
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The financial support of the Taiwan National Science Council is gratefully acknowledged. We also thank Prof. Y.W. Yang and Dr. Jyh-Fu Lee of the Taiwan Synchrotron Radiation Research Center for their help in the extended X-ray absorption fine structure experiments and Ms. Ru-Rong Wu of the National Cheng Kung University for her solid-state nuclear magnetic resonance experimental assistance.
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