Groundwater and Soil Remediation Using Electrical Fields - ACS

Chapter 26

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Groundwater and Soil Remediation Using Electrical Fields 1

2

2

Jiann-Long Chen , Souhail R. Al-Abed , James A . R y a n , 2

and Zhenbin Li 1

2

Department of C i v i l and Environmental Engineering, North Carolina Agriculture and Technical State University, Greensboro, N C 27411 National Risk Management Research Laboratory, U.S. Environmental Protection Agency, Cincinnati, OH 45268

Enhancement of contaminant removal and degradation in low permeability soils by electrical fields are achieved by the processes of electrical heating, electrokinetics, and electrochemical reactions. Electrical heating increases soil temperature resulting in the increase of contaminant volatility and diffusivity and decrease of partitioning coefficients. The main electrokinetic phenomena include electroosmosis and electromigration. Electroosmosis facilitates groundwater flow and transport of nonionic contaminants in low permeability regions while electromigration increases the removal efficiency of ionic contaminants from the regions. Electrochemical reactions utilize high and low redox potentials at the anode and the cathode to oxidize or reduce contaminants, such as trichloroethylene (TCE), to less toxic forms. This article demonstrates the feasibility of reducing T C E to chloromethane at a granular graphite cathode. The results warrant further research to identify the degradation pathways and mechanisms.

434

© 2002 American Chemical Society In Chemicals in the Environment; Lipnick, Robert L., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2002.

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Introduction Remediation of contaminated groundwater and soil has been a major challenge for engineers and scientists since the passages of the Comprehensive Environmental Response, Compensation, and Liability Act ( C E R C L A ) in 1980 and revised Resource Conservation and Recovery Act (RCRA) in 1984. The predominant method for remediating groundwater and soil is pump-and-treat (P&T). P & T operates by pumping contaminated groundwater from an array of wells, removing the contaminants from the water with on-site treatment systems, and re-injecting the treated water back into the aquifer or disposing of it off-site. Despite the popular use of P & T as the treatment method, the available data shows that P & T is inefficient and sometimes unable to restore the groundwater quality to acceptable levels (7-5). The ineffectiveness of P & T on groundwater and soil remediation can be attributed to the presence of low permeability soil, non-uniform flow patterns, and sorption of contaminants. The presence of low permeability soil causes low water recovery rate and requires more pumping power. This results in indefinite pumping periods from wells and excessive operational cost. Non-uniform flow patterns are due to the inter-bedding of soils with high and low permeability, the presence of fracture, or other heterogeneities. When water flows through heterogeneous media, the flow rate in the high permeability is much greater than in the low permeability region. As a result, advective transport, which is the major mechanism for contaminant transport in a P & T system, in the low permeability soil is insignificant. Instead, the dominant transport mechanism of contaminants in these regions is diffusion (/). Sorption processes occur when a solute is bonded to the surface of soil particles either by adsorption, chemisorption, or ion exchange (4). The mass of contaminants sorbed is proportional to the surface area of the soil particles. Since the surface area to mass ratio of low permeability soils, such as clay, is orders of magnitude greater than high permeability soils (5), greater amounts of contaminants may accumulate in the low permeability regions. Thus, during P & T operations, the low permeability regions become reservoirs of contaminants, which are slowly released to the flushing flow by diffusion and desorption mechanisms. The combined effects of non-uniform flow patterns and sorption greatly reduce the amount of contaminants removed and prolongs the remediation period (3,6). Typically, contaminant concentrations rapidly decrease at the beginning of P & T operations as dissolved contaminants are easily removed. After the initial removal period, the contaminant concentration reaches a level in which further pumping is unable to remove the contaminant because of diffusion.

In Chemicals in the Environment; Lipnick, Robert L., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2002.

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The drawbacks of P & T has prompted the development of new treatment technologies that are capable of addressing the shortcomings of P&T. The application of electrical fields is one of many methods applied to facilitate the removal of contaminants (3). When an electrical field is applied to a clay-rich aquifer, the mobilization and removal of contaminants can be accomplished by electrical heating, electrokinetics, and electrochemical degradation. These processes are described in details in the following sections.

Electrical Heating When current flows through a medium, heat energy is generated by resistive heating and causes soil temperature to rise. The magnitude of heat energy generated is a function of the soil electrical conductivity and the gradient of the electrical field (7,8) and can be written as Eq.(l).

Q = opvf

Ο)

where Q = heat energy generated, σ = electrical conductivity (S/m), and V = electrical field (V). Electrical heating usually is combined with other remediation methods, such as soil vapor extraction, air sparging, or P&T, to mobilize and facilitate the transport of contaminants in the low permeability regions. B y heating and thus, increasing the temperature of soils with electrical fields, the removal efficiency of contaminants may be increased. First, the volatility of volatile organic compounds increases with temperature (7,9). This would greatly increase the efficiency of techniques, such as soil vapor extraction, that extract contaminants from the gaseous phase. Secondly, the diffusion coefficient of organic contaminants increase with temperature (9,10). Therefore, heating soil, especially the low permeability zones, can increase the mass transfer rate of contaminants from the diffusion-controlled regions to the advection-controlled regions. Finally, the sorption of neutral organic compounds such as trichloroethylene (TCE) and pyrene, on natural particles decrease with temperature (9,10). As a result of electrical heating, more of the sorbed contaminants partition into the aqueous phase and can be removed by the flushing flow. Thus, the retardation factors of contaminants are reduced and the removal efficiency is increased. It should be emphasized that although other thermal-based remediation methods, such as steam injection, offer some advantages as electrical heating, they suffer from the same difficulties as P & T when low permeability soil is present as the non-uniform flow paths limit the pathways of the steam (11,12). Electrical heating, however, is not limited by the presence of low permeability



437 soil. In fact, the efficiency of electrical heating benefits from the presence of clay minerals that are major constituents in low permeability regions (5). The reason for this is clay minerals increase the electrical conductivity of soil by providing additional paths for current flow through the mineral surface and the double layer, which contains excess cations (13). The low permeability regions, thus, are preferential paths for current flow and the heat generated (Eq. 1) is greater than in high permeability regions. The ability to increase the temperature of the low permeability regions by electrical current has been recognized and applied to enhance contaminant removal efficiency. Buettner and Daily (7) combined electrical heating with air stripping to remove T C E from an aquifer that contained a clay layer. The target volume (89 m ) was heated with six cylindrical electrodes (3.05 m in length and 10.2 cm in diameter). The electrodes were equally spaced and powered by a three-phase source. After 9,600 kW-h of energy input, the T C E concentration in the exit air flow was decreased from 70 ppm to below 1.5 ppm . Due to difficulties in sampling and insufficient monitoring points, the effect of electrical heating on the T C E removal efficiency and mechanisms were not characterized in this study. They estimated that approximately 50% of the energy was applied to the target volume due to the setup of the electrodes. Other limiting factors include the potential hazard of electrical shock to the operator, which resulted in periodical rather than continuous operation of the process. In a more controlled laboratory setup, Heron et al. (9) used electrical heating to enhance the T C E removal efficiency by soil vapor extraction. The authors showed that the steady-state T C E flux in the exit gas was increased from 0.13 to 0.35 g/d when the soil was heated from 23 to 85 °C. The removal rate was further increased to 2.5 g/d when the soil was heated to approximately 100 °C. A t this temperature, however, steam was generated and the column started to dry out, which would cause significant decrease in electrical conductivity i f the soil moisture was not replenished. A t the end of the test, 99.8% of T C E was removed. The enhancement of T C E removal efficiency by electrical heating was attributed to the changes in Henry's constant (H), partitioning coefficient (K ), and gaseous diffusion coefficient (D). These changes are summarized in Table 1. Although electrical heating has been proven to be a feasible method in mobilizing and transporting volatile and semi-volatile contaminants in low permeability soils, sorption of contaminants by the soil particles can still cause the tailing of effluent concentrations to proceed for a long period of time. Even by heating soil to around 100 °C, more T C E was sorbed than dissolved (9). Therefore, sorption/desorption processes, which are least affected by temperature (Table 1), could be the limiting mechanisms for mobilizing contaminants by electrical heating. 3

v

v

d


438 Table 1 Physical Property changes of T C E with Soil Temperature Soil Temperature °C

Henry's constant ratio

(flux/flux23) 1 2.65 18.8

23 85 95 Downloaded by UNIV OF CALIFORNIA SAN DIEGO on January 23, 2016 | http://pubs.acs.org Publication Date: January 24, 2002 | doi: 10.1021/bk-2002-0806.ch026

Observed TCE flux ratio

Partitioning coejf. ratio (Kd 23/Kd) f

(H/H )

(D/D 3) 1 1.33 1.41

23

1 6.8 9.0

Gaseous diffusion coeff ratio 2

1 1.15 1.18

SOURCE: Adapted with permission from reference 9. Copyright 1998 American Chemical Society.

Electrokinetics Electrokinetic remediation of groundwater and soil usually involves the use of electroosmosis and electromigration. Electroosmosis describes the phenomenon of water flow induced by application of an electrical field on porous media (14). Electromigration describes the transport of ions under the influence of an electrical field (75). The electroosmotic flux, q , can be written as eo

Qeo = V OT eo

(2)

e

where v = electroosmotic velocity, r = tortuosity, and n = effective porosity. The electroosmotic velocity can be described by the Helmholtz-Smoluchowski equation (14). eo

e

v„=-^VF

(3)

μ where ε = permittivity of the pore fluid, ζ= zeta potential, and μ = viscosity of the pore fluid. According to Eq.(3), the electroosmotic velocity is directly proportional to the zeta potential and the electrical gradient, and inversely proportional to the pore fluid viscosity. It also indicates that unlike hydraulic flow, electroosmotic flow is independent of pore size (16). Thus, theoretically the electroosmotic flow is unaffected by the presence of low permeability soil with small pore sizes. As clay minerals that possess high zeta potential and electrical


439


conductivity are preferentially found in low permeability regions, the electroosmotic flow tends to be greater than in the high permeability regions. This gives electroosmosis the advantage over conventional methods, such as P&T, in mobilizing and transporting contaminants in the low permeability regions. Cations move from the anode toward the cathode, whereas anions move from the cathode to the anode under the influence of an electrical field. The migration velocity of an ion, v/, in a dilute solution is (77) v,. = M , W

(4)

where u\ = ionic mobility of ion i. The ionic mobility ,defined as the average velocity per unit electrical field strength, is determined from (77)

2

where A[ = molar ionic conductivity (S m /mol), z\ = the valence of the ion, and F = the Faraday constant (96,490 C/mol). Eq.(4) and (5) suggest that the migrating velocity of ions is independent of pore size. As mentioned above, the electrical gradient in low permeability regions is greater than in high permeability regions due to the high clay contents. As a result, ions migrate faster in low permeability than in high permeability regions. Electrokinetic technique has been applied to remove inorganic and organic contaminants from low permeability soils. It should be pointed out that the major mechanisms in removing inorganic contaminants (ionic) are electroosmosis and electromigration, whereas electroosmosis is the main mechanism for removing organic contaminants (non-ionic). Laboratory-scale experiments showed that electrokinetics can removed 75 to 92% of metals, such as C d , Pb , and Z n (18-20). Acar et al. (27) showed 85 to 95% of phenol was removed using electrokinetics in a test column and the retardation was minimal. In another study, Bruell et al. (22) observed 7 to 25% removal of petroleum hydrocarbons and T C E by electroosmosis after 2 to 25 days of processing. Al-Abed and Chen (23) showed more than 90% of T C E mobilization by electroosmosis after 28 days of processing. In larger pilot- and field- scale applications, electrokinetics is combined with other in situ treatment methods to form a treatment system called Lasagna® process (24). The setup in a Lasagna® process includes placing a sheet-like electrodes on each end of the target volume and creating in situ treatment zones between the electrodes. The designing concept is to mobilize and transport contaminants by electrokinetics toward the treatment zones, where they can be degraded. The configuration of the electrodes and treatment zones can be either vertical or horizontal (Fig. 1). Pilot-scale tests conducted by Ho et al. (24) using 2+

2+

2+



440

Vertical Configuration

Degradation Zone

contaminated Degradation soil Zone

Mate: electroosmoticflowis reversed upon switching electrical polarity.

Figure 1. Horizontal and Vertical Configurations of the Lasagna® Process. Reproduced with permission from reference 24. Copyright 1997 Elsevier Science Β. V.


441 a vertical Lasagna® process showed 98% of paranitrophenol (PNP) was mobilized after 0.94 pore volume of electroosmotic flow. After additional 1.6 pore volume of electroosmotic flow, PNP was non-detect (less than 10 ppb) in the soil. Additional field-scale tests (vertical Lasagna® process) by Ho et al. (25,26) achieved T C E removal of 95 to 99% by electroosmosis. They also reported an estimated cost of $45 to $80/yd for remediating soil between depths of 15 to 45 ft by electrokinetics. The horizontal Lasagna® process offers advantages that 1) can be applied at greater depths than the vertical Lasagna® process; and 2) utilizes both electroosmotic and hydraulic flows to remove contaminants. However, the application of such setup has been limited to uncontaminated aquifers (27). Reports that document the removal of contaminants by using horizontal Lasagna® process is unavailable at this time. Researchers at the US E P A are currently conducting a field-scale test using the horizontal Lasagna® process to remove T C E at Offutt A F B near Omaha, N E (28). The results from this study should provide significant performance data for the removal of contaminants with the horizontal Lasagna® process. The first reported field tests on removing heavy metals from groundwater and soil using electrokinetics were conducted by Lageman (29,30). The contaminants included Pb, Cu, Zn, Cd, and As. The tests showed average removal efficiency of 74% for Cu and Pb, 50% for A s and 20% for Zn. It should be noted that out the removal percentages were variable across the test site due to non-uniform electrical fields. Pilot-scale tests conducted by Acar and Alshawabkeh (31) showed that 90% of Zn was mobilized by electrokinetics and 75% of Zn precipitated within 2 cm of the cathode compartment. Precipitation of metals close to the cathode was also observed by Hicks and Tondorf (20) and Eykholt and Daniel (32). This was attributed to the low solubility of metals at alkaline conditions close to cathode during electrokinetic remediation. The precipitation of metals at cathode can be prevented by rinsing away hydroxyl ions generated at the cathode (20).


3

Electrochemical degradation Electrochemical degradation utilizes the high redox potential at an anode and the low redox potential at a cathode to oxidize or reduce contaminants. Most applications of electrochemical degradation involved cathodic reduction or anodic oxidation of chlorinated organic contaminants. Schmal et al. (33) showed that pentachlorophenol was sequentially dechlorinated (i.e. reduced) at a carbon fiber cathode. Experiments conducted by Cheng et al. (34) indicated that the composition of the cathode was critical for cathodic reduction of 4-



442 chlorophenol. Reduction of 4-chlorophenol was not observed when a carbon cloth or graphite rod were used as cathodes. Palladized carbon cloth cathode was shown to dechlorinate 4-chlorophenol to phenol. They hypothesized three possible mechanisms for reduction of 4-chlorophenol to phenol: 1) direct reduction at the cathode, 2) hydrogénation (i.e. reduced by hydrogen gas) at the palladium catalyst surface, or 3) adsorption of 4-chlorophenol at the cathode surface followed by hydrogénation at the palladium/electrode surface. Anodic oxidation was also shown to be a feasible technique for chlorophenol degradation (35). The mechanisms for this process is not well understood. It is hypothesized that chlorophenol radical cations are first formed and subsequently deprotonated. The product undergoes further oxidation to a benzoquinone derivative and ring-opening to acids, such as muconic, naleic, and oxalic and finally to C 0 (36-38). Just like cathodic reduction, the efficiency of anodic oxidation of chlorophenol depends on the material of the electrodes. Generally, oxide-based anodes perform better than metals as they are less prone to the formation of oligomers, which cause the inactivation of the electrodes (35,38,39). 2

Electrochemical Degradation of T C E In an experiment of removing T C E from a natural soil with electrokinetics,. Al-Abed and Chen (23) observed dichloroethylene (DCE) isomers at the vicinity of the cathode. The mechanism of T C E reduction is unclear, however, the low redox potential and ample electrons available at the cathode suggest that T C E might be reduced at the surface of the graphite cathode. To our best knowledge, cathodic reduction of TCE, although theoretically possible, has not been experimentally proven. The authors have conducted experiments to prove the occurrence of cathodic reduction of T C E using graphite cathodes. The rest of this article describes their approaches and results.

Experimental Setup and Approach The setup included two major chambers: a reactor where cathodic reduction occurred and a solution reservoir. The reactor was a glass cylinder 15 cm in length and 5 cm in diameter. The reactor was filled with 10 cm of granular graphite (30-50 mesh) as the anode and 4 cm of granular graphite as the cathode. A 1-cm layer of silicate sand separated the anode and cathode. The solution reservoir was a 500 mL gas-tight glass cylinder. The reactor and reservoir were hydraulically connected with a peristaltic pump and Teflon tubing. The whole system was kept gas-tight during operation to prevent loss of



443 T C E and its degradation products. The solution matrix was 400 m L of 0.1 M ammonium acetate with 250 mg/L of sodium azide. The addition of sodium azide was to prevent the T C E from being degraded by microbial activities. The solution was put inside the reservoir, spiked with 2.8 μΐ of pure T C E , and shaken at 100 rpm for at least 12 h. The solution was pumped at the designated flow rate (3.8 mL/min) from the reservoir to anode, through cathode, and back to the reservoir. The T C E concentration was monitored by taking samples from the reservoir. Electrical power was not applied until the T C E concentration reached steady-state, which usually took 12-24 h. Three different voltages (5, 10 and 20 V ) were applied between the anode and cathode to evaluate the effects of voltage on the reduction rate. The durations of the experiments were between 14 and 60 h. The test under 10 V lasted for only 14 h because of an unrepairable leak in the system.

Analytical Methods At each sampling period, 5 m L of solution and 50 μΐ of head space gas from the solution reservoir was extracted and analyzed. The solution was spiked with 50 μι of 1000 ppm benzene-D6 as a internal standard. Afterwards, the sample was shaken in a head space sampler (HP 7694E) at 70 °C for 20 minutes. The concentrations of vinyl chloride, D C E isomers, T C E , and other chlorinated products that could be daughter products of T C E were analyzed with a gas chromatograph (HP 6890) equipped with a mass selective detector (HP 5973). After the analysis of organic compounds, pH was measured and the chloride concentration in the sample was analyzed with a Capillary Ion Analyzer (Waters CIA). The 50 μΐ of head space gas was directly injected into a G C (HP 6890) with a HP Plot/Al 0 column (50m χ 0.53 mm χ 15 μηι) and the concentrations of methane, ethane, ethylene, and acetylene were quantified with a FID detector. 2

3

Results and Discussion T C E monitoring before the initiation of electrical power indicated T C E was sorbed on the granular graphite. Assuming a linear isotherm for the sorption of TCE, we estimated the average partitioning coefficient, K , to be 2.2±0.52 mg/g (n=3). The concentration of T C E decreased with time under all the voltage applied (Fig. 2). The degradation rate of T C E increased with applied voltage. Assuming the reduction of T C E is pseudo-first-order with respect to the aqueous concentration of T C E , the reaction rate constants determined from Fig. 2 at applied voltages of 5, 10, and 20 V are 0.01, 0.04, and 0.06 h" , respectively. The half-life, t , of T C E at applied voltages of 5, 10, and 20 V are 69, 17, and 12 h, respectively. Thus, increasing applied voltage from 5 to 20 V decreases the half-life of T C E by more than 80%. d

1

50



444 The major product of T C E reduction was chloromethane (CM) and no other chlorinated daughter products were observed. This is unexpected as reduction of T C E usually results in the formation of D C E isomers, which are further reduced to vinyl chloride and to ethane or ethylene (40). The pathway of C M formation from T C E reduction is unclear. Judging from the fact that C M was the only chlorinated product, we hypothesize that T C E was sorbed on the graphite surface at the cathode and both the carbon-carbon and carbon-chlorine bonds were cleaved. Additional studies are needed to determine whether carbon-carbon bond is cleaved before carbon-chlorine bond, or vise versa.

υ

30

40

Time (h)

Figure 2. Disappearance of TCE with Time at Applied Voltages of 5 (circle), 10 (square), and 20 V (triangle).

The rate of C M generation is proportional to the applied voltages (Fig. 3). The mole of C M generated in Fig. 3 includes C M that is in the aqueous phase and in the head space. To estimate the amount of C M in the head space, we used Henry's Law constant of 9.55 Latm/mole for C M (10). The amount of C M sorbed on the granular graphite was not included since the K for C M was unavailable. Therefore, the amount of C M generated shown in Fig. 3 is an underestimation of the total amount of C M in the system. At 5V, the moles C M during the test were less than the moles T C E added. At applied voltages of 10 and 20 V , however, the moles C M generated were larger than the moles T C E added after 10 h of operation (Fig. 3). This suggested that for each mole of T C E degraded, more than one mole of C M was generated at 10 and 20 V of applied voltages. Thus, this supports our hypothesis that the carbon-carbon bond of T C E is cleaved. The products underwent further reactions to form C M . The ratio of moles C M generated to moles of T C E degraded was estimated to between 2 and 3. A ratio of 2 would d



445 suggest that the carbon in C M comes only from TCE, whereas a ratio of more than 2 suggests the carbon in C M might come from other sources, such as graphite or acetate. Graphite or acetate as the sources of carbon were confirmed in several control experiments. The solutions in controls contained no T C E , however, methane and ethane formations were observed when an electrical field was applied. The maximum ratio is 3 as each mole of T C E contains 3 moles of chlorine. Further studies that quantify the amount of C M sorbed on granular graphite are needed to determine the moles of C M generated per mole of T C E degraded. In addition to the formation of C M , chloride concentration increased with operation time. With the measurements of T C E , C M , and chloride concentrations in the solution and head space, it is possible to conduct a mass balance of the test based on the total mass of chloride. The procedures are as the following: 1) Calculate the mass of chloride in initial T C E ; 2) Calculate the mass of chloride in T C E ; 3) Calculate the mass of chloride in C M ; and 4) Subtract the results of 2) and 3) from the result of 1). The resulting mass of chloride was converted to concentration and taken as the estimated chloride produced. It should be noted that the estimation assumed no sorption of C M on the graphite. The estimated chloride production was compared to the measured chloride concentration (Fig. 4). The results indicate the estimated and measured chloride concentrations agree with each other. This agreement also suggests that sorption of C M on graphite was minimum and hence, can be neglected. pH is one of the important solution parameters that changes as a result of an applied electrical field. The change is due to electrolysis at the anode and cathode. Protons and oxygen are produced at the anode, whereas hydroxyl ions and hydrogen are produced at the cathode. The pH with time at all applied voltages had similar patterns. It increased from p H 6.7 to 7.7 shortly after the initiation of power (Fig. 5). The increase of pH was due to hydroxyl ions that were pumped from the cathode compartment into the solution chamber. The magnitude of p H jump increased with the applied voltage. This is because the rate of hydroxyl ion production increased with applied voltage. After 6 to 7 h of power application, the pH started decreasing with time suggesting protons generated at the anode finally reached the solution chamber (Fig. 5). One of the objectives of this study was to degrade T C E to non-toxic compounds, such as ethane and chloride, using cathodic reduction. This was not achieved because most of the product was C M . However, long term application of electrical field might change the compositions of the degradation products as C M could undergo further reduction once T C E is completely depleted from the system. Experiments are being conducted at US E P A , Cincinnati, O H to investigate the long-term effects of electrical field on the compositions of T C E degradation products.


446 1.8 1.6

ο Ε J

• •

• 1.4

•

5 Volt 10 Volt 20 Volt

1.2 1.0 0.8


0.6 0.4 • 0.2

I*

0.01 10

20

30

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Time (h)

Figure 3. Ratio of mole CM generated to the initial mole of TCE as a function of time at applied voltages of 5 (circle), 10 (square), and 20 V (triangle).

0.5

1.0

1.5

2.0

2.5

3.0

3.5

4.0

C f Produced-Estimated (ppm)

Figure 4. Comparison of Estimated Chloride with Measured Chloride Production at an Applied Voltage of 5 V.

References 1. 2 3.

Mackay, D . M ; Cherry, J.A. Environ. Sci. Technol. 1989, 23, 630-636. MacDonald, J.A.; Kavanaugh, M . C . Environ. Sci. Technol. 1994, 28, 362368. N R C , Groundwater and Soil Clean Up: Improving Management of Persistent Contaminants; National Academy Press: Washington, DC, 1999, 285pp.


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Figure 5. pH of Solution with Time at Applied Voltages of 5 (circle), 10 (square), and 20 V (triangle).

4. 5. 6. 7. 8. 9. 10.

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