Environ. Sci. Technol. 2000, 34, 1535-1541
Cyclodextrin-Enhanced Electrokinetic Removal of Phenanthrene from a Model Clay Soil SEOK-OH KO,† M A R K A . S C H L A U T M A N , * ,‡, # A N D ELIZABETH R. CARRAWAY# Environmental Research Team, Daewoo Institute of Construction Technology, Seoul, Korea, Department of Agricultural and Biological Engineering, Clemson University, Clemson, South Carolina 29634-0357, and Department of Environmental Toxicology and Clemson Institute of Environmental Toxicology, Clemson University, Pendleton, South Carolina 29670
Removal of hydrophobic organic contaminants (HOCs) from saturated low-permeability subsurface environments using a solubility-enhanced electrokinetic remediation process is demonstrated for a model system. Phenanthrene, hydroxypropyl-β-cyclodextrin (HPCD), and kaolinite were selected as a representative HOC, HOC solubility-enhancing agent, and model clay soil, respectively. Electrokinetic (EK) column experiments were conducted under various operating conditions, and the results were interpreted in terms of the EK properties and expected phenanthrene solubilization of the test systems. No significant effects of HPCD on the EK properties of kaolinite were observed. Initial pore solution pH values dictated the initial electroosmotic flow (EOF) and charge flow rates through the test samples. However, with increasing EK operating times, low pH values (i.e., near or below the point of zero charge of kaolinite) dominated over most of the column length in unbuffered systems, thereby decreasing the EOF and charge flow rates with time. To minimize these adverse effects, pH control of the anode reservoir with a Na2CO3 buffer was used to keep EOF and charge flow rates high. EK experiments using HPCD solutions showed greater phenanthrene removal from the kaolinite samples, and the removal efficiency depended on the HPCD concentration used. Longer EK operating times without pH control were generally not beneficial for removing phenanthrene because of the low EOF rates obtained after 3 days. The best overall phenanthrene removal was obtained by flushing the anode reservoir with a high HPCD concentration prepared in the Na2CO3 buffer solution. The results obtained from this preliminary study show that an EK process combined with HPCD flushing and pH buffering may be a good remediation alternative for removing HOCs from lowpermeability subsurface environments.
Introduction The widespread presence of hydrophobic organic contaminants (HOCs) in subsurface environments has prompted * Corresponding author phone: (864)656-3250; fax: (864)656-0338; e-mail:
[email protected]. † Daewoo Institute of Construction Technology. ‡ Department of Agricultural and Biological Engineering. # Department of Environmental Toxicology and Clemson Institute of Environmental Toxicology. 10.1021/es990223t CCC: $19.00 Published on Web 03/08/2000
2000 American Chemical Society
rigorous development of in-situ remediation technologies for sites contaminated with these compounds. Because of their low solubilities and slow dissolution/desorption rates, HOCs are typically associated with solid matrices and thus are difficult to remove from subsurface environments using traditional pump-and-treat technologies. Although alternative methods to water-based flushing (e.g., bioremediation, solvent washing, surfactant-enhanced aquifer remediation) have been developed for relatively permeable subsurface systems contaminated with HOCs, these technologies often are not directly applicable to fine-grain soils and sediments because of the low permeabilities and resistance to hydraulic flow through these types of porous media. Electrokinetic (EK) remediation is an emerging technology that can effectively remove metals and soluble organic pollutants from saturated fine-grain soils (1-3). As an insitu process, EK remediation has two major advantages over pump-and-treat systems: (1) a high level of control over flow direction and (2) the ability to work in low-permeability clays and/or silt-laden soils (1, 2). Several research sponsors, including the U.S. Department of Energy (DOE), U.S. Environmental Protection Agency (EPA) and the Electric Power Research Institute (EPRI), have been funding studies of various EK remediation techniques for use in site cleanups (2). In general, highly soluble inorganic compounds including alkali metals, chlorides, and nitrates are easily removed from low permeability subsurface environments in EK applications, and heavy metals such as lead, mercury, cadmium, and chromium also have responded favorably, some with the aid of complexing agents (4-6). For example, Pamukcu and Wittle (7) reported that the transient acid front movement from the anode region toward the cathode is beneficial for metal desorption/dissolution and contributes to the removal process. Generally, analysis of the EK removal of heavy metals is based on the temporal and spatial variation of species because the speciation and solubility of most heavy metals depend greatly on the pH and redox conditions present; once the temporal and spatial pH and redox variations are obtained, metal speciation can then be calculated from equilibrium pC-pH-p relationships (8). During EK operation, a reduction in the remediation efficiency due to metal precipitation is possible in a soil because of the transport of hydroxide ions from the cathode reservoir, where they are generated, toward the anode. Several methods have been proposed to minimize this problem. For example, Lageman et al. (9) controlled undesirable chemical reactions by filling the electrode reservoirs with chemical buffer solutions and obtained an average removal of 74% for lead and copper. Hicks and Tondorf (8) rinsed the cathode with water to keep pH values lower and observed a high removal rate for zinc. Yeung et al. (10) injected the watersoluble chelating agent ethylenediaminetetraacetate (EDTA) into electrode reservoirs to keep lead from precipitating above pH 5. Li et al. (11) installed a cation selective membrane in front of the cathode to prevent the hydroxide ions from advancing toward the anode. Organic compounds having high water solubilities (i.e., low sorption/distribution coefficients) tend to orient similarly to water molecules in an applied electric field and move toward an electrode depending on their charge, polarity, and mobility and on the surface charge characteristics of the saturated soil medium. Removal of these compounds generally results from the combined effects of the electroosmotic flow (EOF) of water and the electromigration (EM) of dissolved ions toward the appropriate electrode. In some cases, competition between EOF and EM may occur, and the VOL. 34, NO. 8, 2000 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
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transport becomes highly dependent on pH which in turn depends on the electrode reactions. For example, Shapiro and Probstein (1) examined the removal of phenol and acetic acid from a clay and obtained efficiencies of 75-95% and 42-95%, respectively, under various EK conditions. They also observed a change in the ζ potential of the clay from negative to positive that resulted in lower EOFs and corresponding lower removal efficiencies. In some experiments, Shapiro and Probstein observed that the EOF direction was toward the anode instead of the more-commonly observed flow pattern toward the cathode, a phenomenon that has also been observed in some other EK studies (10, 12). Based on their results, Shapiro and Probstein recommended the use of suitable pH buffering agents to ensure the long-term viability of EK processes. Although EK remediation results to date have shown excellent removal efficiencies for metals and soluble organic contaminants from low permeability subsurface environments, problems with removing HOCs still exist because of their low aqueous solubilities and high preferential sorption to soils and aquifer materials (1). For example, Bruell et al. (13) conducted column tests to evaluate the effectiveness of EOF in mobilizing dissolved hydrocarbons and trichloroethylene (TCE) from a saturated clay and found that the compounds having the highest distribution coefficients (e.g., hexane, isooctane) showed lower removal rates than the other compounds examined (e.g., benzene, toluene, TCE). Therefore, to effectively utilize EK processes in HOC remediation applications, methodologies to increase HOC mobility may need to be incorporated into EK operations. The objective of this study was to make a preliminary evaluation of a relatively new process for removing HOCs from contaminated low-permeability subsurface systems. The process, which we have termed solubility-enhanced EK remediation, has the potential to overcome many of the obstacles facing HOC removal from fine-grain soils by integrating the beneficial effects of enhanced HOC solubilization by surfactants or other appropriate facilitating agents with the accelerated transport (e.g., EOF) resulting from EK processes, thereby accelerating the overall HOC mass transport and subsequent rate of HOC cleanup. For all potential solubility-enhanced EK remediation applications, selection of a suitable HOC facilitating agent is a critical step. For example, surfactant sorption on finegrain soils could conceivably alter the EK response of the soil particles and potentially reduce the EOF. Likewise, appreciable surfactant sorption to fine-grain soils can create a new immobile partitioning compartment that will decrease the overall effectiveness of such a process (14-16). Ideally, the optimum facilitating agent will have a high affinity for enhancing HOC solubilities and a low potential for sorption to the solid phase and/or pore blockage; unfortunately, these two performance criteria often conflict with one another. For example, results from our previous studies (14-17) suggest that conventional micelle-forming chemical surfactants (e.g., sodium dodecyl sulfate, Tween 80) may not be suitable for solubility-enhanced EK remediation applications because although they are excellent solubility-enhancing materials, they show appreciable sorption to clay mineral surfaces. Conversely, cyclodextrins (CDs) that have been modified to increase their solubility in water may be more appropriate facilitating agents because of their low affinity for sorption to the solid phase and their reasonably satisfactory HOC solubilization capabilities (e.g., ref 17 and references therein).
Experimental Methods Information on the experimental materials used in this study and many of the analytical procedures have been previously reported (14, 16-18). Briefly, phenanthrene (distilled water 1536
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FIGURE 1. Schematic diagram of electrokinetic cell (not to scale). solubility of 1.28 mg/L (14)) was chosen as a representative HOC, and hydroxypropyl-β-cyclodextrin (HPCD) was selected as the facilitating agent for phenanthrene removal from a saturated model clay soil. HPCD, a cyclic oligosaccharide formed from the enzymatic degradation of starch by bacteria, increases the apparent aqueous solubility of many HOCs (e.g., complexation constant between phenanthrene and HPCD of ∼3000 to 3600 M-1 (17)) and exhibits negligible sorption onto fine-grain soils (17), characteristics ideal for effective remediation applications as discussed above. A schematic of the EK test cell constructed for this study is shown in Figure 1. A glass column (10 × 2.5 cm, Ace Glass Co.) was utilized to minimize phenanthrene sorption to the cell wall. The column was equipped with several ports to measure the electric potential at various locations within the soil column. Platinum was used to fabricate the active electrodes; although more costly than the graphite electrodes often used in EK work, the latter electrodes have high HOC sorption capacities that would have adversely affected the experiments. Constant voltage across the cell was applied with a DC power supply (Tektronix CPS250) and current was monitored with an ammeter (Elenco M-1000B). The voltage drop across test samples was measured at the passive electrodes using a digital multimeter (Tektronix DM254). The test column was connected to electrode reservoirs containing the platinum anode and cathode; entry of clay particles into the reservoirs was prevented by placing glass filter discs at the junctions. Scaled standpipes connected to ports in the electrode reservoirs were used to vent the oxygen and hydrogen gases generated from the anode and cathode, respectively, and to measure the EOF; pressure induced flow was minimized by adjusting the static head in the standpipes. Additional details of the experimental setup of the EK columns are provided in Ko (18). Sample Preparation. EK experiments were conducted using kaolinite (Sigma) as a model low-permeability subsurface material. To simulate a site having fine-grain soils contaminated with a high sorbed HOC concentration and to obtain a uniformly distributed loading on the kaolinite for systematic evaluation of the HOC removal efficiencies, phenanthrene contaminated samples were prepared using the following procedure. Initially, an aqueous phenanthrene solution just below the solubility limit was prepared in 0.01 M NaCl. The phenanthrene solution was added to air-dried kaolinite to obtain a solid-to-solution mass ratio of 1:3.3 (i.e., 300 g/L) and then thoroughly mixed on a shaker table for 2 days. After equilibration, the suspension was centrifuged at 5000 rpm for 5 min, the supernatant was decanted, and the kaolinite was resuspended once more in the phenanthrene solution and mixed for 5 days. This procedure was repeated
one additional time so that the resulting kaolinite suspension would have a high initial concentration of phenanthrene in both the aqueous and solid phase. Upon final equilibration, the suspension was centrifuged at 9000 rpm for 30 min, and the pH of the supernatant was measured. The water content of the centrifuged kaolinite was ∼45 to 50%, thereby allowing direct packing of the column with the centrifuged samples. To fill the EK test column with the prepared kaolinite samples, the column was positioned vertically, and a fraction of the sample was added followed by tapping the cell on a table to remove air bubbles. Another fraction was then added, and the procedure was repeated until the appropriate soil column length was obtained. After the column was filled, a subsample of the remaining kaolinite was extracted as described below to determine the total initial phenanthrene present in the column (nominally 1.9 mg/kg-kaolinite); although this procedure did not allow for direct phenanthrene quantification in the individual aqueous and solid phases, calculations based on measured solid-to-water ratios and a distribution coefficient of 0.002 L/g (14) showed that over 80% of the phenanthrene initially present in the column was adsorbed to the kaolinite. Permeability tests using aqueous solutions with and without HPCD were conducted to determine if blockage of the kaolinite pores by the facilitating agent would occur in our systems; in these studies, a falling-head method suitable for samples of low permeability (19) was used directly with the packed soil columns. Other important physical parameters for the kaolinite samples such as bulk density and porosity were determined using standard volume-weight relationships (19); additional details are provided in Ko (18). Data Acquisition. After filling the anode and cathode reservoirs with appropriate flushing solutions, a total potential of 5 or 14 V was applied to the test column such that the initial voltage drop across the kaolinite sample was 1.3 or 5 V, respectively, as measured with the passive electrodes installed in the sample. During an experiment, the current through the cell and voltage drop across the sample were both monitored with time. Changes in the head in the standpipes were measured with time to estimate the EOF. For experimental systems utilizing HPCD flushing solutions, the anode reservoir was filled with a solution containing 1.37 or 6.85 mM HPCD, and the EK experiments were run using the same procedures as described above; based on measured complexation constants (17), the phenanthrene distribution between HPCD and water for these two HPCD concentrations ranged from 5-6:1 and 21-26:1, respectively, depending on the total phenanthrene concentration. At the end of each run, the aqueous solutions in the anode and cathode reservoirs were collected for pH and phenanthrene concentration measurements. The kaolinite test sample was extruded and sectioned, and pH and phenanthrene measurements were made on subsamples from each section. For pH measurements, a fraction of each kaolinite section was added to a 50 mL polyethylene centrifuge tube and centrifuged at 9000 rpm for 30 min to elute the pore solution; the pH of the pore solution was then measured using a microcombination pH electrode capable of measuring the pH in a sample as small as 10 µL. For phenanthrene analysis, another small fraction of each section was added to a 25 mL glass centrifuge tube (Corex). Methanol was then added to the tubes before placing them on a shaker table for 7 days to extract the phenanthrene from the solid phase. After centrifuging the samples at 7000 rpm for 20 min, phenanthrene concentrations in the supernatant were determined using a spectrofluorometer (Photon Technology International) and appropriate external standard solutions. The water content of each kaolinite section was determined by comparing the weight of the remaining clay section before and after drying in an oven at 105 °C for 24 h.
TABLE 1. Experimental Conditions for Electrokinetic (EK) Tests EK exp. no.
voltagea (V)
1 2 3
5 (1.3) 14 (5) 14 (5)
4 5 6
14 (5) 14 (5) 14 (5)
purging solutionb NaCl NaCl NaCl for the first 4 days, followed by Na2CO3 for the rest of the test 1.37 mM HPCD; NaCl 6.85 mM HPCD; NaCl 6.85 mM HPCD; Na2CO3
duration (days)
initial pH
∼6 ∼6 ∼8
6.0 6.0 4.0
∼14 ∼6 ∼6
4.8 4.0 4.0
a Total applied voltage. The values in parentheses were the applied voltage across the kaolinite samples. b All solutions had a total ionic strength of 0.01 M fixed by the electrolyte shown (i.e., 0.01 M NaCl or 0.0033 M Na2CO3).
EK Experimental Conditions. Parameters associated with each experiment are listed in Table 1. A background electrolyte of 0.01 M NaCl or 0.0033 M Na2CO3 was added to the electrode reservoirs to maintain a fixed ionic strength of 0.01 M; to further aid pH control in the buffered experiments, a fresh 0.0033 M Na2CO3 solution was continuously fed into the anode reservoir at a rate of 0.25 mL/min using a peristaltic pump (Cole-Parmer Model 7553-70) to help neutralize the protons generated at the anode. Results from purging solutions with and without HPCD were compared to evaluate its ability to enhance the removal of phenanthrene from the contaminated kaolinite.
Results and Discussion Typical physical parameters for the kaolinite samples tested were as follows: bulk density of 1.26 g/cm3, porosity of 0.53, and permeability of 1.36 × 10-10 cm2 for a 0.01 M NaCl aqueous solution. The permeability of a kaolinite sample measured using a 6.85 mM HPCD solution was approximately 1.31 × 10-10 cm2; this insignificant difference implies that the HPCD molecules were able to pass through the kaolinite pores without creating any blockages. This is a critical test that all potential facilitating agents must pass before they can be considered further for use in fine-grain subsurface systems; for HPCD, passing this first test meant that its effectiveness as a facilitating agent for HOCs warranted further examination. It should be noted that the above values for the porosity and, particularly, the permeability of our samples were higher than those commonly measured for real clay soils. This difference can be attributed directly to the procedures we utilized to contaminate the kaolinite and pack the test columns. Because our principal objective was to evaluate the effectiveness of a solubility-enhanced EK remediation process for a low-permeability material contaminated with an HOC, it was more important for us to obtain well-known and uniform starting conditions than to closely replicate the physical properties of an actual field site. Thus, it must be remembered that although our experimental results can be compared directly to one another for relative effects, they cannot be expected to predict the performance of remediation at a real field site. Electrokinetic Characteristics of Kaolinite. (A) pH Distribution. It is well-known that the EK properties of soils (e.g., ζ potential) are directly related to pore solution pH conditions (1, 16). Measured pH values in the kaolinite sections after completion of EK experimental runs are shown in Figure 2a for flushing solutions that did not utilize HPCD. For Experiments 1 and 2, pore solution pH values were very acidic (pH ∼2) near the anode and very alkaline (pH ∼12) near the cathode after completion of the runs; the low and high pH conditions that developed in the anode and cathode regions, respectively, resulted from the electrolysis reactions VOL. 34, NO. 8, 2000 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
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FIGURE 2. Distribution of pH in the saturated kaolinite samples at the end of each EK experimental run. Normalized distance from the anode represents the normalized length of kaolinite in the EK cell. See Table 1 for further explanation of experimental conditions. (a) Experiments 1 and 2: NaCl solution. Experiment 3: NaCl solution for 4 days followed by Na2CO3 buffer solution. (b) Experiments 4 and 5: NaCl solution containing 1.37 and 6.85 mM HPCD, respectively. Experiment 6: Na2CO3 buffer solution containing 6.85 mM HPCD. of water at the electrodes. Along the length of the kaolinite samples, the relative transport of protons toward the cathode and hydroxide ions toward the anode, the buffering capacity of the samples, and the water dissociation reaction determine the final pH distribution observed. The higher mobility of the proton relative to the hydroxide ion (20, 21) leads to the condition where low pH values dominate throughout most of the kaolinite sample (i.e., from the anode up to a normalized distance of ∼0.7); at a normalized distance greater than 0.7 from the anode, Experiments 1 and 2 showed a sharp increase in pH up to the cathode region. Conversely, the pH distribution at the end of Experiment 3 showed a much different profile than the first two experiments; for the latter experiment, purging was conducted for 4 days with the NaCl solution before the Na2CO3 buffer solution was fed into the anode reservoir as described in the Experimental Section. As shown in Figure 2a, the buffer solution effectively maintained a high pH (∼9) in the anode region. Before flushing the anode reservoir with the buffer solution, the kaolinite sample likely had a similar pH distribution as Experiments 1 and 2. When the Na2CO3 buffer solution was used for purging, however, the pH throughout the kaolinite sample gradually increased until at the end of the run it had reached a value g 6 over most of the column. Experiment 3 indicates that using an appropriate buffer solution during an EK run can maintain pore solution pH values at higher levels, a condition that is more favorable for creating appreciable EOF rates through the kaolinite (see below). pH distributions for the kaolinite samples flushed with HPCD solutions are shown in Figure 2b. For HPCD solutions prepared in 0.01 M NaCl (Experiments 4 and 5), pH distributions at the end of the runs were similar to those not using HPCD, indicating that the facilitating agent did not influence the pH; the slightly lower pH values observed at the end of Experiments 4 and 5 (with HPCD) versus 1 and 1538
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FIGURE 3. Charge flow rate expressed as the average electrical conductivity (σe) during each EK run. See Table 1 for further explanation of experimental conditions. (a) Experiments 1 and 2: NaCl solution. Experiment 3: NaCl solution for 4 days followed by Na2CO3 buffer solution. (b) Experiments 4 and 5: NaCl solution containing 1.37 and 6.85 mM HPCD, respectively. Experiment 6: Na2CO3 buffer solution containing 6.85 mM HPCD. 2 (without HPCD) resulted from the lower initial pH condition (Table 1). Additionally, the slightly lower anode region pH observed for Experiment 4 (1.37 mM HPCD) versus that for Experiment 5 (6.85 mM HPCD) likely resulted from the longer duration of the run (i.e., 14 versus 6 days, Table 1). Higher pH values were observed throughout the kaolinite at the end of the experimental run that used 6.85 mM HPCD prepared in the Na2CO3 buffer (Experiment 6, Figure 2b). Although the initial pore solution of the kaolinite was slightly acidic (pH 4), the carbonate and bicarbonate ions effectively neutralized the protons generated during the run such that at the end of 6 days pH values throughout the sample were g 7. (B) Charge Flow. The average electrical conductivity (σe, units of [A/m2]/[V/m]) of a saturated clay soil characterizes the charge flow through the soil during an EK application (22). Under initial EK operating conditions, σe is simply the conductivity of the soil pore solution which can be calculated from theoretical considerations. However, because the chemical composition of the pore fluid changes temporally and spatially with EK operation, a spatially averaged electrical conductivity based on adherence to Ohm’s law is often used instead. In this study, σe was estimated by dividing the interstitial current density, which is the movement of charge per unit area per unit time, by the voltage drop across the kaolinite sample (22). Values for σe for a saturated soil are typically in the range of about 0.01-1.0 (A/m2)/(V/m) and depend on several properties of the soil, including the porosity, degree of saturation, and composition of the pore solution (22). Values for σe calculated as described above are shown in Figure 3a for Experiments 1-3. A comparison between the σe values for Experiments 1 and 2 show lower values for the former because of the lower applied voltage used (Table 1), in agreement with Ohm’s law. Note that the charge transport as characterized by σe is directly related to the movement of water by electroosmosis (i.e., EOF) when the flow is characterized by the average electroosmotic permeability (18);
therefore, both parameters will show similar behaviors with EK operating time (data not shown). In Experiments 1 and 2, σe decreased with time as the EK properties of kaolinite changed; in particular, as the low pH front developed throughout the sample (Figure 2a), the ζ potential of kaolinite decreased resulting in a reduction of water flow as well as charge flow. Initial σe values from Experiment 3 showed a similar decreasing trend with time as was observed in Experiment 2; however, when the Na2CO3 buffer solution was used to purge the anode reservoir after 4 days, σe values began to increase because of the increasing pH which, in turn, created an increasingly negative kaolinite ζ potential. The average electrical conductivity for EK experiments using HPCD is shown in Figure 3b. Compared to the values observed in the absence of HPCD, it can be seen that there was no significant influence of the HPCD concentration on σe for samples flushed with the NaCl background solution. As for Experiment 2, the electrical conductivity in Experiments 4 and 5 gradually decreased with time because of the predominance of low pH across the kaolinite sample which resulted in a decrease in charge flow. When the Na2CO3 buffer solution was used (Experiment 6), the charge flow also showed an initial decreasing pattern with time before becoming constant after about 4 days; it is not known whether σe values would have remained constant for longer operating times. For example, Dzenitis (22) conducted an EK experiment using a concentrated NaOH solution to keep the anode reservoir pH above 9; the electrical conductivity in his study showed an initially quick drop before maintaining a constant value for about 3 days followed by an eventual rise again. Unfortunately, the short experimental operating time used in the present study does not provide insight into the longterm electrical conductivity trend that will prevail when using a Na2CO3 buffer solution. Future studies with extended operating times will be necessary to gain this type of longterm operating information. However, it is evident that charge transport will be greater in EK systems that use buffers to maintain optimum pH conditions. (C) Electroosmotic Flow (EOF). For negatively charged saturated soils, movement of the positively charged counterions in the diffuse double layer toward the cathode creates an EOF of pore water during EK processes. The departing positive ions in the diffuse layer are immediately replaced by other positive ions supplied from the bulk aqueous solution, thereby providing a constant restoration of charge equilibrium in the double layer and continuous EOF. This is critical in the removal of uncharged, relatively nonpolar organic contaminants from saturated clay soils because, in general, removal of these compounds will be highly dependent on the EOF rate generated during an EK process (1-3, 6, 13). For the EK experiments conducted using only a 0.01 M NaCl aqueous solution in the anode and cathode reservoirs (Experiments 1 and 2), EOF rates decreased with increasing operating time (Figure 4a); from the cumulative curves shown in Figure 4, EOF rates at any time can be estimated from the slopes of the curves. The cumulative EOF transported to the cathode region during Experiment 2 was much greater than Experiment 1 which can be attributed to the higher voltage applied in that run (Table 1); note that all other conditions (e.g., initial pH) were otherwise the same for these two experiments. According to Ohm’s law, a higher applied voltage results in an increased electric current that represents the transport of ions when a conductive medium has the same resistivity; this faster ion transport results in the faster transport of water by electroosmosis as discussed above. EOF rates decreased with time because the EK properties of the kaolinite sample (e.g., pH, σe) were changing with time as shown previously. For example, the EOF rate for Experiment 1 varied from about 5.1 mL/day (superficial velocity of
FIGURE 4. Cumulative electroosmotic flow (EOF) collected during each EK run. See Table 1 for further explanation of experimental conditions. (a) Experiments 1 and 2: NaCl solution. Experiment 3: NaCl solution for 4 days followed by Na2CO3 buffer solution. (b) Experiments 4 and 5: NaCl solution containing 1.37 and 6.85 mM HPCD, respectively. Experiment 6: Na2CO3 buffer solution containing 6.85 mM HPCD. 1.4 cm/day) at initial times to 2.13 mL/day (superficial velocity of 0.6 cm/day) at the end of the test. For Experiment 2, the EOF rate was about 13.6 mL/day (3.9 cm/day) at initial times and ended at about 4.9 mL/day (1.3 cm/day). Note that because the EOF rate gradually decreases with time for these two experiments, removal of an uncharged nonpolar organic contaminant will not be directly proportional to the overall time of the EK operation. Experiment 3 was run under similar conditions as Experiment 2, except that a Na2CO3 buffer solution was used as a purging solution after about 4 days of operation (Table 1); additionally, a lower initial soil solution pH was used in Experiment 3 (i.e., pH 4 versus 6). Similar to the previous experiments, the cumulative EOF for Experiment 3 gradually decreased with time when the NaCl solution was present in the anode reservoir. After switching to the buffer solution, however, the EOF began to increase indicating that the buffer ions were effective in bringing the pH back up to a level that resulted in a more efficient EK operation. It is well-known that the pH of a soil solution has a significant effect on the ζ potential of a clay (22, 23); the ζ potential of the kaolinite used in this study is zero at approximately pH 4.2 (i.e., point of zero charge, PZC, of 4.2) and becomes increasingly negative with increasing pH (16). According to the HelmholtzSmoluchowski equation, the EOF in a saturated clay soil should be proportional to the ζ potential; therefore, faster EOF rates toward the cathode are expected for the kaolinite used here when the pH becomes increasingly higher than 4.2. Similar pH effects are also likely the reason that the initial EOF rates of Experiment 3 were lower than those of Experiment 2 even though both experiments used the same applied voltage. However, it is noteworthy that the net EOF in all of our experiments was always toward the cathode, even when pore solution pH values dropped well below the PZC in select experiments. Whether this observation is related to the inherent complexity of clay minerals, the relative affinity of water molecules for cations versus anions, or some other phenomenon is not presently known. VOL. 34, NO. 8, 2000 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
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FIGURE 5. Phenanthrene removal efficiencies from the saturated kaolinite samples at the end of selected EK runs. See Table 1 for further explanation of experimental conditions. Experiments 2, 4 and 5: NaCl solution containing 0.00, 1.37, and 6.85 mM HPCD, respectively. Experiment 6: Na2CO3 buffer solution containing 6.85 mM HPCD. Cumulative EOFs for EK experiments using HPCD prepared in NaCl or Na2CO3 solutions are shown in Figure 4b. When EK experiments were conducted with HPCD prepared in NaCl, cumulative EOFs showed similar trends to those experiments without HPCD present. For example, Experiment 4 (1.37 mM HPCD) had an EOF rate that decreased from an initial value of 13.5 mL/day (superficial velocity of 3.9 cm/day) to 0.9 mL/day (0.25 cm/day) at the end of test (14 days). Likewise, in Experiment 5 (6.85 mM HPCD) the rate changed from 12.1 mL/day (3.5 cm/day) initially to 3.8 mL/day (1.1 cm/day) after 6 days. The difference between Experiments 4 and 5 with respect to the cumulative EOF volumes collected during the first 6 days likely results from the different starting pH conditions as discussed previously. Of interest is the large decrease in the EOF rate (i.e., lower slope) observed for the longer EK run (i.e., 14 days for Experiment 4); this result is consistent with the pH distribution shown in Figure 2b where lower pH values were obtained in Experiment 4. The cumulative EOF for the HPCD solution prepared in the Na2CO3 buffer (Experiment 6) continuously increased as the experiment progressed (Figure 4b). For this experiment, the EOF rate increased from an initial value of 4.1 mL/day (superficial velocity of 1.2 cm/day) to 26.1 mL/day (7.2 cm/ day) at the end of the 6-day run. Again, this result corresponds well with the pH distribution in the kaolinite sample when the buffer solution was used (Figure 2b); because the pH was maintained at much higher values by the buffer, a higher negative kaolinite ζ potential was maintained which led to higher EOF rates. Electrokinetic Removal of Phenanthrene. Phenanthrene removal efficiencies at fixed locations in the EK test columns were calculated from the total (i.e., adsorbed plus dissolved) initial phenanthrene mass present (nominally 1.9 mg/kgkaolinite) and the mass remaining in extruded and sectioned samples at the end of each run. Results from select experiments are shown in Figure 5. When no buffer was used for pH control, the total eluted volume of solution passing through the clay was approximately 75.6, 70.3, and 40.6 cm3 for Experiment 2 (no HPCD), Experiment 4 (1.37 mM HPCD), and Experiment 5 (6.85 mM HPCD), respectively. Based on an estimated total pore volume of about 15.6 cm3 for the kaolinite samples, these cumulative EOFs correspond to approximately 4.85, 4.51, and 2.6 pore volumes (PVs), respectively. Although the cumulative EOF was greatest for the solution without HPCD (Experiment 2), phenanthrene removal averaged only about 25%; for comparison, the 6.85 1540
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mM HPCD purging solution (Experiment 5) showed approximately a 56% removal for the same EK operating time (∼6 days). These results confirm that HPCD enhances phenanthrene desorption from kaolinite, in agreement with predictions made by Ko et al. (17). Also, it is worth noting that although Experiment 4 (1.37 mM HPCD) lasted for 14 days and had nearly two times greater cumulative EOF than did Experiment 5 (6.85 mM HPCD), phenanthrene removal turned out to be much lower for the former experiment (33% versus 56% removal). Thus, it can be concluded that HOC removal is dependent on the HPCD concentration. Also, it can be concluded that the 14 day run for Experiment 4 did not greatly increase phenanthrene removal because the EOF rate was very low after 3 days (Figure 4b). This last result implies that an EK operation using either water purging or an HPCD flushing solution will be less effective if the pH of the saturated soil is not controlled. This is demonstrated convincingly in Experiment 6, where phenanthrene removal was approximately 75% for a 6.85 mM HPCD solution prepared in the Na2CO3 buffer (Figure 5). In this experiment, the cumulative EOF volume flushed through the kaolinite was approximately 117.3 cm3 (7.52 PVs) after 6.2 days, a volume much greater than that observed for the other experiments that operated for a similar time period (i.e., ∼6 days for Experiments 2 and 5). Correspondingly, the higher average phenanthrene removal observed in Experiment 6 can be attributed to the combined effects of both beneficial processes (i.e., enhanced desorption/solubility of phenanthrene by HPCD and increased EOF rates due to maintaining higher pH values).
Acknowledgments We gratefully acknowledge the constructive comments of the three reviewers and their suggestions for improving the paper. Preliminary funding for this work was provided by the Office of the Vice President for Research and Associate Provost for Graduate Studies, through the Center for Energy and Mineral Resources, Texas A&M University. Additional funding was provided by the Gulf Coast Hazardous Substance Research Center, which is supported under cooperative agreement R822721-01-4 with the United States Environmental Protection Agency. The contents of this paper do not necessarily reflect the views and policies of the U.S. EPA nor does the mention of trade names or commercial products constitute endorsement or recommendation for use.
Literature Cited (1) Shapiro A. P.; Probstein R. F. Environ. Sci. Technol. 1993, 27, 283-291. (2) Shannon D. Environ. Sci. Technol. 1995, 29, 452A. (3) Ho SA. V.; Sheridan P. W.; Athmer C. J.; Heitkamp M. A.; Brackin J. M.; Weber D.; Brodsky P. H. Environ. Sci. Technol. 1995, 29, 2528-2534. (4) Acar, Y. B.; Alshawabkeh, A. N.; Gale, R. J. Waste Management 1993, 13, 141-151. (5) Acar, Y. B.; Alshawabkeh, A. N. Environ. Sci. Technol. 1993, 27, 2638-2647. (6) Trombly, J. Environ. Sci. Technol. 1994, 28, 289A-291A. (7) Pamukcu S.; Wittle J. K. Environ. Prog. 1992, 11, 241-250. (8) Hicks R. E.; Tondorf S. Environ. Sci. Technol. 1994, 28, 22032210. (9) Lageman R.; Pool W.; Seffinga G. Chem. Ind. 1989, 18, 587-590. (10) Yeung, A. T.; Hsu, C.-N.; Menon, R. M. J. Geotech. Eng., ASCE 1996, 122, 666-673. (11) Li, A.; Yu, J.-W.; Neretnieks, I. Environ. Sci. Technol. 1998, 32, 394-397. (12) Eykholt, G. R.; Daniel, D. E. J. Geotech. Eng., ASCE 1994, 120, 797-815. (13) Bruell C. J.; Segall B. A.; Walsh M. T. J. Envir. Eng., ASCE 1992, 108, 68-83. (14) Ko, S.-O.; Schlautman, M. A.; Carraway, E. R. Environ. Sci. Technol. 1998, 32, 2769-2775.
(15) Ko, S.-O.; Schlautman M. A. Environ. Sci. Technol. 1998, 32, 2776-2781 (16) Ko, S.-O.; Schlautman, M. A.; Carraway E. R. Environ. Sci. Technol. 1998, 32, 3542-3548. (17) Ko, S.-O.; Schlautman, M. A.; Carraway E. R. Environ. Sci. Technol. 1999, 33, 2765-2770. (18) Ko, S.-O. Ph.D. dissertation, Texas A&M University, College Station, TX, 1998. (19) Methods of Soil Analysis; Part 1, American Society of Agronomy: Madison, WI, 1965. (20) Jacobs, R. A.; Sengun, M. Z.; Hicks, R. E.; Probstein R. F. J. Environ.
Sci. Health 1994, A29, 1933-1955. (21) Alshawabkeh, A. N.; Acar, Y. B. J. Environ. Sci. Health 1992, A27, 1835-1861. (22) Dzenitis, J. M. Environ. Sci. Technol. 1997, 31, 1191-1197. (23) Hunter, R. J. Zeta Potential Colloid Science; Academic Press: New York, 1981.
Received for review February 24, 1999. Revised manuscript received September 30, 1999. Accepted January 11, 2000. ES990223T
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