© 2001 by CRC Press LLC Chapter Three Separation Techniques © 2001 by CRC Press LLC 3.1 In Situ Remediation of Contaminated Soils by Electrokinetic Processes Sibel Pamukcu Department of Civil and Environmental Engineering Lehigh University Bethlehem, Pennsylvania C.P. Huang Department of Civil and Environmental Engineering University of Delaware Newark, Delaware Introduction Soil systems are subject to contamination by a variety of hazardous chemicals, such as heavy metals and toxic organic compounds. The major sources of pollutants are attributed to landfills and industrial operations. For example, any hazardous substance present in a soil matrix represents a threat to public health and groundwater. The latter is one of the most valuable natural resources and a major source of drinking water in the United States. Many domestic, industrial, and agricul- tural activities depend on groundwater resources. Therefore, strategies for soil clean-up are increas- ing in demand. Most of the host of soil remediation techniques available commercially are subject to a variety of restrictions during application. Ex situ treatments such as pump and treat and containment can be costly and therefore not totally attractive. Techniques including bioremediation, vitrification, freezing, and soil washing are some of the options available, but they are usually very site specific and do not offer a good prospect of in situ treatment. Vitrification and freezing do not extract contaminants from soils and therefore cannot be considered ultimate clean-up options. Bioremediation is limited by a number of technical difficulties such as nutrient transport and acclimation of microorganisms, among others. Few contaminants can be effectively removed by soil washing. Accounting for all of these obstacles, there is a necessity to develop new alternatives for in situ soil clean-up. Electrokinetic processes treatment has emerged as a potential technique for in situ decontamination of contaminated soils. This is the same process used previously by geological engineers to consolidate foundations for construction. Electrokinetic treatment is an in situ treatment process that is capable of simultaneously transporting inorganic and organic compounds in porous media, including radionuclides. The electrokinetically aided transport is based on well-known electrokinetic processes primarily com- posed of electroosmosis, electrophoresis, and ion migration in wet soil. The two primary mechanisms © 2001 by CRC Press LLC that mobilize contaminants are (1) the movement of the charged species by electromigration or electro- pheresis; and (2) the transport of contaminants by the advection of electroosmotic flow. The rate and efficacy of these processes are dependent on the type of soil and contamination. The treatment involves applying a low direct current (on the order of milliamps per square centimeter of the cross-sectional area of the electrodes), or a low potential gradient (on the order of a few volts per centimeter) between electrodes inserted in the soil. As a result, the contaminants are transported toward the anode or cathode electrode sites by ionic or electrophoretic migration, and electroosmotic advection. The contaminants are then removed at the electrode sites by one of several different methods. These methods include electroplating, adsorption onto the electrode, precipitation or co-precipitation at the electrode, pumping near the electrode, complexing with ion exchange resins, or capturing in reactive permeable barriers. While electroosmosis is analogous to soil washing, electromigration is the primary mechanism of electrokinetic transport when the contaminants are ionic or surface charged (Acar et al., 1989; 1990; Pamukcu and Wittle, 1992a; b; Probstein and Hicks, 1993; Reddy and Parupudi, 1997). Past experience with electrokinetic process in contaminated porous media has shown that the process is most effective when the transported substances are dissolved in the pore fluid, surfaces charged, or in the form of small micelles with little drag resistance (Electorowicz and Boeva, 1996; Hamed et al., 1991; Pamukcu and Wittle, 1992a; 1993a; b; Pamukcu et al., 1995b; 1997; Pamukcu, 1994; Pamukcu and Pervizpour, 1998). Background Overview Research in electrochemical treatment for the purpose of restoring contaminated subsurfaces has accel- erated in the past two decades. Some of the currently researched methods of electrochemical treatment are referred to as electrokinetic extraction, electrokinetic barriers, electrobioremediation, electrostabili- zation (injection), and electrocontainment. Earlier work in the mid-1970s and early 1980s focused on utilizing the technique for soil densification to improve performance of containment facilities. Later, studies focused on the effects of electrolysis soil chemistry and the use of electrokinetics for contaminant removal from soils. Most of this work was conducted on the laboratory scale and some on the pilot scale. The first field study was published in 1988 (Banerjee et al., 1988) as a feasibility study of potential application of electrokinetics for chromium removal from subsurfaces. Research in 1989 first showed the importance of the process-generated pH gradients between anode and cathode. In the same year, field applications attempted to alleviate the effects of the pH gradients by controlling the chemical environment around the electrodes. In 1991, the effects of speciation and precip- itation on the efficiency of electrokinetic transport of metal ions through soil were presented. Since the early 1990s, numerous laboratory studies have substantiated the applicability of the technique to a wide range of contaminants in soils. Among the contaminants shown to react to electrochemical treatment in the laboratory and some in the field are non-aqueous phase liquids such as chlorinated hydrocarbons, mononuclear aromatic hydrocarbons (MAHs), polynuclear aromatic hydrocarbons (PAHs), phenols, sul- furous, and nitrogenous compounds, and heavy metals. More recently, integrated methods of soil restoration that rely on electrochemical technology as well as other technologies (e.g., bioremediation, funnel-and-gate, and reactive membranes) have been introduced; and some have been demonstrated in the field, such as the Lasagna Soil Remediation project (Ho et al., 1995). Furthermore, powerful analytical models and their numerical solutions have been developed; this has helped to better understand the underlying mechanisms of transport of single and multiple ionic species under constant or transient electric fields. These laboratory studies have clearly shown that electrochemical treatment is a powerful in-situ process that can be used to simultaneously treat inorganic and organic compounds in porous media. However, the technology must be used judiciously in the field because each contaminated site is unique. The application must be engineered to site specifics and the treatment steps must be sequenced properly for an optimum solution. Soils are heterogeneous, silty, and contain fine metallic oxide and colloidal organic and inorganic © 2001 by CRC Press LLC substances. In field situations, the contaminants are often found adsorbed onto soil surfaces, iron oxide coatings, soil colloids, and natural organic matter, or retained in clay interstices as hydroxycarbonate complexes, or in the form of immobile precipitates in soil pore throats and pore pockets. It is now well recognized that the contaminated soil becomes dynamically complex under an applied electrical potential. The solid and liquid components of the soil are reactive, which allows complex electrochemical reactions to take place. Given such conditions, it may be preferable to base the treatment on a phenomenological approach using site-specific information rather than on analytical models of well-controlled systems. Historical Development In 1808, Reuss observed electrokinetic phenomena when a dc current was applied to a clay-water mixture. Water moved through the capillary toward the cathode under the electric field. When the electric potential was removed, the flow of water immediately stopped. Napier (1846) distinguished electroosmosis from electrolysis; and in 1861, Quincke found that the electric potential difference through a membrane resulted from streaming potential. Helmholtz was the first to treat electroosmotic phenomena analytically in 1879. A mathematical basis was provided by his work. Pellat (1904) and Smoluchowski (1921) later modified it to apply to electroosmotic velocity. Out of this treatment of the subject, the well-known Helmholtz-Smoluchowski (H-S) theory was developed. The H-S theory deals with the electroosmotic velocity of a fluid of certain viscosity and dielectric constant through a surface-charged porous medium of electrokinetic potential (zeta, ζ ), under an electric gradient. The H-S equation is: (3.1.1) where, u eo = Electroosmotic velocity ε = Dielectric constant of pore fluid ζ = Zeta potential of soil particles µ = Viscosity of fluid ∂φ / ∂ x = Electric gradient (field strength) It must be noted that Eq. (3.1.1) is valid only for large pores in which the electrical double layer is small compared with the pore radius, and all the mobile charge is assumed to be concentrated near the pore wall. In 1939, Casagrande demonstrated that applying electro-osmosis to soils with high water content caused such an increase in the effective stress that the gain in shear strength kept steep slope cuts remain stable. Casagrande indicated that small reductions in water content by electroosmosis could produce significant increases in soil strength. Since then, electrochemical treatment of soils has been investigated and used in many field projects, including improvement of excavation stability, electrochemical hardening, stabilization of fine-grained soils, consolidation, and densification. In the late 1960s and early 1970s, direct current was successfully applied to recover residual oil from deep-seated geological formations (Enhanced Oil Recovery) (Waxman and Smits, 1967; Amba et al., 1964). Utilization of direct current to drive contaminants out of the soil pores started in the late 1970s and early 1980s. Segall and co-workers reported detection of high concentrations of metals and organic compounds in electroosmotically drained water of a dredged sludge in the field in 1980. Since then, successful applications of the electrochemical decontamination technique have been demonstrated on pure soil-contaminant mixtures in the laboratory by numerous researchers (Pamukcu et al., 1990; 1995; 1997; Hamed et al., 1991; Bruell et al., 1992; Acar et al., 1992; 1994; 1995; Probstein and Hicks, 1993; Runnels and Wahli, 1993; Ugaz et al., 1994; Hicks and Tondorf, 1994; Eykholt and Daniel, 1994; Pamukcu, 1994; Yeung et al., 1996; Alshawabkeh and Acar, 1996a; b; Dzenitis, 1997). Theoretical Aspects Electrokinetic phenomena in a porous medium are based on the relative motion between a charged surface and the bulk solution at its interface (Adamson, 1986; Hunter, 1981). The formation of an electric u eo ες µ φ∂ x∂ = © 2001 by CRC Press LLC double layer at the charged surface of clay particles explains these electrokinetic phenomena of interest: electroosmosis, electrophoresis, and electromigration. The Electric Double Layer Consider a negatively charged clay particle surface in contact with a water solution of ions. The attraction of counter ions and repulsion of co-ions, when combined with the diffusion along concentration gradients and the mixing by random thermal motion of the ions, leads to the formation of an electric double layer (Gouy, 1910; Chapman, 1913). According to Stern (1924), the electric double layer is composed of a fixed layer (Stern layer) and a diffuse layer (Gouy layer). In the Stern layer, the ions are assumed to oscillate about fixed adsorption sites, whereas in the diffuse layer, ions are assumed to undergo Brownian motion. In a porous plug of clay, the surface becomes negatively charged when wetted with water. This charge is balanced by the adjoining Stern and Gouy layers, which carry the positively charged ions. The thickness of the Stern layer is approximately the radius of a hydrated cation adsorbed on the clay particle surface. The Stern and Gouy layers are divided by three planes: one is the plane of the claywater interface; the second is the outer Helmholtz plane (OHP); and the third is the plane of shear. The OHP is the plane that defines the outer limit of the Stern layer, the layer of positively charged ions condensed onto the clay particle surface. The drop in potential in the Stern layer is linear from the surface potential of ψ o to ψ d at the OHP. The plane of shear is the plane at which the mobile portion of the diffuse layer can slip or flow past the charged surface. The potential at this shear plane is referred to as the electrokinetic potential, or zeta ( ζ ) potential. The potential distribution in the diffuse layer is given by the Poisson-Boltzmann equation, which describes an exponential fall of the potential. ψ x = ψ o e  κ x (3.1.2) where κ = Reciprocal thickness of the diffuse double layer Ψ x = Potential at distance x from the OHP or surface Ψ o = Potential at the OHP or surface Integrating the Poisson-Boltzmann equation with appropriate boundary conditions will provide the thickness of the diffuse layer, which is indirectly related to the ionic concentration in the bulk solution and the valence of the counter-ions. (3.1.3) where: e =Electron charge k = Boltzmann constant z i =Ionic charge or valence n i× = Ionic concentration in the bulk solution Electrophoresis Electrophoresis is defined as the migration of charged colloids in a solid-liquid mixture under an electric potential gradient. This migration is the movement of colloidal particles, not small ions. For clay-water systems, if we place a direct current (dc) field across its suspension, negatively charged clay particles migrate toward the anode. The unrestrained particle transport through water in a poorly consolidated system will likely compact the soil to the anode and disintegrate it on the cathode side. In a compact system of a porous plug, electrophoresis is of less importance due to the restrained solid phase. But in the process of soil decontamination under direct current, electrophoresis of clay colloids could still play an important role if the migrating colloids have the toxic chemicals adsorbed onto them. This was demonstrated by Grolimund et al. (1996), who showed strongly sorbing lead was transported by mobile colloids. κ Qe 2 εkT   z 2 i n i ∞ i ∑ 12/ = © 2001 by CRC Press LLC An important contribution of electrophoretic movement to contaminant transport may be when the contaminants are in the form of colloidal electrolytes or ionic micelles. Micelle formation is promoted as the concentration of the aggregating groups increases. Ionic micelles often carry a high charge and exhibit high conductance in dilution. The conductance increases with increasing concentration owing to buildup of charge with further aggregation. However, at a critical concentration, a sudden and sharp decrease in conductance occurs that is attributed to (1) increasing association of the ionic colloids, which results in increased fraction of neutral colloids; and (2) retarding inter-ionic forces. Evidence of micellar transport was observed in a study by Pamukcu (1994), in which highly mobile anionic-surfactant micelles facilitated the transport of nonpolar organic compounds toward the anode in the opposing direction of electroosmotic flow. Electroosmosis Electroosmosis is the complement of electrophoresis. The latter involves discrete particle transport through water, while electroosmosis is the transport of water through a continuous soil particle network. The diffuse layer of water close to the solid surface contains an abundance of counter charges (cations) to balance the surface charge deficiency. These counter charges are strongly held on the surface and diffuse away toward the free water in the middle of the pore. The section referred to as free constitutes the pore water that is free to flow under a hydraulic gradient. When an electric field is applied, the surface or particle stays fixed, while the mobile diffuse layer moves, carrying the adjacent water with it. The fluid on the surface is set into motion due to the electromigration of the cations contained in it. As the cations start shearing toward the negative electrode, the thick fluid of the surface layer is dragged along. The velocity of this motion is zero at the solid surface and maximum at the plane of shear, which can slip or flow past the charged surface. This interface velocity sets the central or free pore fluid in motion. It is not clear how the central portion moves, but it is usually assumed to be viscous drag. The water molecules, being slightly positive because of dipolar fluctuations, may also contribute to the movement of the central layer. The liquid transport in porous media by a combination of these processes is known as electroosmosis. In negatively charged clay particles, an abundance of cations in the diffuse layer generate a net water flow toward the negative electrode (cathode). The ability of electroosmosis to produce a rapid flow of water in a compact, low-permeability soil makes it a significant contributor to soil decontamination processes by advection. Inside the soup of dissolved, suspended, and particulate matter residing in the pore space, the charged species are expected to move independently through the fluid as long as there is connectivity of the fluid phase. The others are carried or advanced to the next locale by the electroosmotic flow of the fluid. During electroosmosis, diffuse layer charges are displaced and polarized in the direction of flow, thus producing a potential difference between the electrode locations. This effect is called the streaming potential, which may decrease the effect of electroosmosis by reversing the polarity in the soil. Electroosmotic flow was shown to be independent of pore size distribution or the presence of macropores (Acar and Alshawabkeh, 1993). Therefore, electroosmosis may be an efficient method to generate a uniform fluid and mass transport in clayey soils. The relative contributions of electroosmosis and ion migration to the total mass transport vary according to soil type, water content, types of ion species, pore fluid concentration of ions, and processing conditions. Electroosmotic advection is most useful for transporting contaminants in clays and low permeability soils because the electroosmotic conductivities of clays are often several orders of magnitude higher than their hydraulic conductivities. Electroosmotic advection is able to transport nonionic and nonpolar as well as ionic species through soil pores toward the cathode. This is best achieved when the state of the material (dissolved, suspended, emulsified, etc.) is suitable for the flowing water to carry it through the tight pores of soil without causing an immovable plug of concentrated material to accumulate at some distance from the electrode. In 1952, Schmid presented the following equation to explain the electrokinetic phenomenon in special cases of very small pores, where it is postulated that the cations are uniformly distributed across the pore cross-sectional area (Mitchell, 1970): © 2001 by CRC Press LLC (3.1.4) In this equation, r is the pore radius, q is the volume charge density, and F the Faraday constant. It is noticed that the flow is independent of the system pore size in the Helmholtz-Smoluchowski equation, while, according to Schmid, the flow depends on the square of the mean pore radius. Meanwhile, neither theory allows for an excess of electrolyte in the pores beyond the number of cations needed to balance the negative surface charge of clay particles. In other words, the influence of the bulk electrolyte concen- tration is neglected. Esrig and Majtenyi, based on the attempt made by Oel in 1955 to unify the two previous theories, have presented a simple equation that appears to include both the Helmholtz-Smoluchowski and Schmid theories (Esrig and Majtenyi, 1966): (3.1.5) where ρ is the average mobile excess electric charge density and d is a parameter characterizing the double layer. According to the authors, this equation can be used with any of the existing double layer theories; it also permits the estimation of fluid velocities for a wide range of capillary sizes. A simplification of Eq. (3.1.5) results in (Casagrande, 1949): (3.1.6a) (3.1.6b) Q eo = k e i e A (3.1.6c) in which Q eo represents the electroosmotic flow rate, k e the coefficient of electroosmotic permeability, i e = V/L the electrical potential gradient, and A the cross-sectional area of flow. The above equation is very similar to Darcys equation for hydraulic flow through a soil column: Q h = k i h A (3.1.7) where i h is the hydraulic gradient, A the cross-sectional area, and k the permeability of the soil. However, the hydraulic and electroosmotic permeability (k and k e , respectively) have different properties. The electroosmotic permeability k e depends primarily on the pore area and is independent of the size of the individual pores; k is very strongly influenced by the actual pore size (Casagrande, 1949). Casagrande (1952) established that k e for almost all soils in which electroosmotic treatment is feasible varies within only about one order of magnitude, with an average value of about 5 × 10 5 cm 2 V 1 s 1 . Thus, estimates of flow rates can be made directly without using any of the kinetic models leading to Eqs. (3.1.6a), (3.1.6b), or (3.1.6c), provided the value of k e , the electroosmotic water flow rate, can be predicted by knowing A and i e . Although Eq. (3.1.6c) describes the theoretical rate of fluid flow in a soil core under potential gradient, there is some uncertainty associated with the effect of interfering factors such as the possible compression of the double layer because of high salinity, loss of electrical energy through the electrolysis of water, changing soil structure, and the reactions of electrode material with the chemicals in water (Ray and u eo r 2 qF 8 µ V L = u eo 1 2 1 d r +   r 2 µ ρ V L ln= u eo Q e A 1 2 1 d r +   r 2 µ ρ ln   V L == Q eo 1 2 1 d r +   r 2 µ ρ ln   V L A= © 2001 by CRC Press LLC Ramsey, 1987). Significant implications of the electrochemistry associated with the electroosmosis process may influence the efficiency of the remediation technique. Spieglers friction model (1958) showed that electroosmotic water transport per unit electrical charge increases with increasing cation: water ratio in the system. Experimental evidence of this theory has been given by a number of researchers (Gray and Mitchell, 1967). An extension of the H-S theory considers a portion of the electric current transported near the surface of or through the solid phase (Wiedemann, 1856). The resulting equation is often referred to as the current efficiency, (time rate of volume of water flow per quantity of electricity), of the system: (3.1.8) where Q eo = Electroosmotic flow rate I =Current r = Radius of the capillary λ 0 = Specific conductance of the bulk liquid λ s = Surface conductance of the capillary wall Surface current is due to the ionic motion in the diffuse layer. In narrow capillaries with low ionic concentrations, thus thick diffuse layers, a disproportionate fraction of the current flows in this layer due to the low conductivity of the bulk fluid. Experimental evidence shows that the current efficiency, Q/I, decreases with increasing ionic concentration in the bulk fluid (Wittle and Pamukcu, 1993). This can be readily explained from Eq. (3.1.2) because ζ and ε / µ are expected to decrease, and λ 0 to increase with increasing ionic concentration of the bulk fluid. The surface conductance also changes with ionic con- centration. As the ionic concentration in the bulk liquid increases, the diffuse double layer shrinks toward the particle surface and the shear plane shifts away from the particle surface so that the majority of the charge is now compensated by the immobile Helmholtz layer. Therefore, the charge density in the diffuse layer decreases, giving rise to a lower surface conductivity, λ s . As a result of this lowered conductivity, a smaller portion of the current flows on the capillary surface. In contrast, in the presence of low ionic concentrations, the diffuse double layer is swollen and much of the charge is compensated by the ions in the diffuse layer. Therefore, the capillary surface conductivity is high and so is the fraction of the current that is transported on the surface. The significance of surface conductance on the prediction of electroosmotic flow as it relates to contaminant migration was investigated by Khan (1991). He proposed a modified theory of electroos- motic velocity of water through soil. In this theory, the true electroosmotic flow is directly proportional to the current carried by the charged solid surfaces in soil. The soil is modeled as parallel resistances of the soil surface and pore fluid, and the zeta potential used in H-S theory is replaced by the surface potential, ψ d , at the Outer Helmholtz Plane (OHP): (3.1.9) where R s = Surface resistance of soil I s = Surface current of soil L =Length With u eo /I s shown to remain fairly constant for clays of different surface conductivity and also pore fluid electrolyte concentrations below 10 2 M, experimentally, Eq. (3.1.3) was further reduced to: Q eo εςI u λ o 2 λ s r +   –= u eo εψ d µ I s R s L = © 2001 by CRC Press LLC u eo = K I s (3.1.10) where K ={ ε ψ d / µ } R s /L = Constant The modified theory basically emphasized that the surface conductivity of the porous compact medium is the most essential precondition for electroosmotic water flow, thus uncoupling it from the water drag component of the migrating ions in pore fluid of high ionic concentration. This theory is in agreement with Spieglers theory of water: cation ratio, as well as Gray and Mitchells (1967) approach of a co-ion exclusion principle based on Donnan theory of membrane equilibrium (1924). Additional evidence to support this finding was presented by Pamukcu and Wittle (1992) for a variety of ion species, where the ionic concentration effect on the measured current efficiency appeared to be most pronounced in clays with high anion retention capacity. At the same concentrations of dilute solutions of electrolytes, kaolinite clay with higher anion retention capacity (poor co-ion exclusion) showed consistently higher electroosmotic flow than montmorillonite clay with lower anion retention capacity (good co-ion exclusion). This obser- vation suggested that the anionic dragging of water toward the anode diminished the net flow toward the cathode compartment in the montmorillonite clay. The zeta potential in Khans (1991) model is defined at the outer limit of the Stern layer and to be a constant surface potential that is invariant with respect to electrolyte concentration. Therefore, the true electroosmotic flow becomes independent of electrolyte concentration in the pore fluid. Results reported by Yin and co-workers (1995) support Khans theory. They found that there is no apparent relationship between electroosmotic mobility and the applied electric field. The term electroosmotic mobility refers to the average velocity achieved by the pore water relative to the solid skeleton, due to an externally applied electrical field of unit strength. The mobility appeared to be proportional to the specific con- ductance of the soil specimen. The mobile ions in the pore solution primarily come from the surface of the clay particles; thus, a higher ionic concentration and hence a higher conductance for clay with a lower initial water content are expected. For kaolinite, Yin et al. (1995) concluded that a mobility value of 0.6 × 10 4 cm 2 /s. volt and a specific conductance of 0.4 m.mho/cm are representative values, and they showed these values do not vary appreciably under low electric field and constant water content. Based on the above discussion, the electroosmotic flow velocity can be expressed as: u eo = K′E (3.1.11) where K ′ = Constant (electroosmotic mobility) E = Electric gradient ( ∂φ / ∂ x) It should be noted that the electroosmotic mobility should not be treated as a phenomenological constant. Electroosmosis velocity can be approximated by the Helmholtz-Smoluchowski equation (Eq. 3.1.1). According to Shapiro and Probstein (1993), for a typical water-saturated clay, with ζ potential of 10 mV, and an electric field strength of 100 V/m, the electroosmotic velocity has a value of 10 6 m/s or ~10 cm/day. Notably, this is at least 10 times lower than the electromigration velocity (Acar and Alsh- wabkeh, 1993). Therefore, in ion-rich pore fluids, the electroosmotic transport of ions becomes negligible compared with electromigration (Baraud et al., 1997). It must be noted that the above derivations are mostly applicable for saturated porous media. Water flow behavior of an unsaturated soil is totally different from that of a saturated system. In the presence of an electrical field, a friction force is created when water molecules begin to move in the soil pores. The frictional stress decreases as the thickness of the water layer increases. For an unsaturated soil- water system, the water layer is extremely thin, usually ranging from 10 10 cm to 10 8 cm. Under such circumstances, all water molecules exhibit strong frictional interaction with the soil surface. In the case of a saturated water-capillary system, the radii of capillaries are relatively large, ranging from 10 1 cm to 10 3 cm. As a result, most capillary water molecules do not interact physically or chemically with the © 2001 by CRC Press LLC capillary wall (Yukawa et al., 1991). Recently, Chang et al. (2000) have proposed a semi-empirical equation for the prediction of electroosmosis flow under unsaturated soil-water system based on the finite plate model. They reported the following expression: (3.1.12) where K is a characterized coefficient (i.e., K = kfw/ µρ 2 Σ 2 ). This characterized coefficient, K, collects several physical properties of the soil-water system such as the fluid density ( ρ ), the specific surface area ( Σ ), the width of the water layer (w), and the fluid viscosity ( µ ). Electromigration Electromigration, or ionmigration, is the primary mechanism of electroremediation when the contam- inants are ionic or surface charged. Speciation and precipitation are major factors in mobilization and transport of heavy metal constituents by the ionmigration component of electrokinetics. The speciation is dependent on a number of fairly well-understood parameters, including pH, redox potential, and ion concentration. These same factors influence the equilibrium conditions relating to both the soil and the contaminants. Charged ions moving toward the oppositely charged electrode relative to solution is called electro- migration. In a dilute system or a porous medium with moderately concentrated aqueous solution of electrolytes, electromigration of ions is the major cause of current conduction. Electromigration velocity measures ion movement in the pore water caused by the electric field at infinitely dilute solutions: (3.1.13) where u m = Electromigration velocity z = Valence or charge of ion F = Faraday constant R = Universal gas constant T = Absolute temperature (K) D * = Effective diffusion coefficient of ion The convective-diffusion equation used to describe the transport of a contaminant through porous media is given by (Shapiro et al., 1989): (3.1.14) where for each species i, c i is the concentration in moles per unit of volume, D i the diffusion coefficient, τ the experimental tortuosity factor, u e,i the electromigration velocity in the x direction, U c , the convection velocity in the x direction, and R i the molar rate of production due to chemical reactions. The electromigration velocity u e,i is represented by: U e,i = U m /τ 2 (3.1.15) where R, T, z i , F, and φ are the gas constant, temperature in units of Kelvin, charge number, Faraday constant and electrical potential, respectively. The convection velocity, or electroosmotic velocity, can be written as: Q eo k' σ e V L   ϖm ρ   2 Kσ e V L   ϖ 2 == u m zF RT D ∗ φ∂ x∂ –= C i ∂ t∂ D i τ 2 ∂ 2 C i ∂x 2 ∂ ∂x C i u ei, u c + ()[] – R i += [...]... Anode A-Cap GS-cap Sample C-Cap Cathode Error (%) Normalized Distance from anode (%) Pb Conc (mg/L) Pb Mass per Section (mg) Mass Fraction (%) 1 20 40 60 80 1 1606.67 576.67 516.67 9 53. 33 1 233 .33 237 6.67 2190.00 1992.27 96.82 211. 83 358.45 5 03. 20 955.42 38 9.82 1.00 0.05 0.11 0.18 0.25 0.48 0.20 27 1606.67 83. 33 3 93. 33 180.00 2 03. 33 266.67 39 6.67 21 83. 33 233 .33 1670. 93 13. 70 1 63. 65 78.12 86.21 132 .80... TABLE 3. 1.1 Percent Removal of Heavy Metals from Clays and Clay Mixtures by Electrokinetic Treatment Soil Typea Metal KS KG KH MS SS As(V) Cd(II) Co(II) Cr(VI) Cs(I) Hg(II) Ni(II) Pb(II) Sr(II) U(V) Zn(II) 54.7 94.6 92.2 93. 1 71.9 26.5 88.4 69.0 97.8 79 .3 54.6 56.8 98.2 93. 9 94.8 80.1 13. 1 95.4 75.2 99.5 84 .3 43. 3 27.2 92.7 95.9 97.6 74.7 42.5 93. 9 66.9 96.0 67.4 36 .3 64 .3 86.6 89.4 93. 5 54.7 — 93. 6... Hazardous Mixed Wastes by Bio-Electrokinetic Remediation and Other Competitive Technologies, In Remediation of Hazardous Waste Contaminated Soils, Eds D L Wise and D J Trantolo, Marcel Dekker, Inc., Ch 18, pp 405– 436 Mitchell, J.K., (1970), In-place treatment of foundation soils: electro-osmosis, J Soil Mech Found Div., ASCE, 96, 92–1 03 Mitchell, J.K and Yeung, T.C., (1991) Electro-Kinetic Flow Barriers... Electrokinetics for removal of low-level radioactivity from soil, 14th Ann DOE Low-Level Radioactive Waste Mgmn.t Conf., Phoenix, Arizona, November, 1 8-2 0, Conf-921 137 -Proc, 256–278 Pamukcu, S and Wittle, J.K., (1993a), Electrokinetically enhanced in situ soil decontamination, in Remediation of Hazardous Waste Contaminated Soils, Wise and Trantolo, Eds., Marcel Dekker, New York, chap 13, 245–298 Pamukcu, S.,... contaminants, and colloidal and micellar particles through low-permeability soils that are not amenable to hydraulic treatments (Acar et al., 1989; 1990; 1991; Acar and Alashawabkeh, 19 93; Alshawabkeh and Acar, 1996a; Bruell et al., 1992; Dzenitis, 1997; Pamukcu and Wittle, 1992; 19 93; Pamukcu et al., 1990; 1991; 1995; Pamukcu, 1994; 1997; Pamukcu and Pervizpour, 1998; Shapiro and Probstein, 19 93; Wittle and. .. hydrocarbons (PAHs), phenols, sulfurous and nitrogenous compounds, and heavy metals (Acar et al., 1986; 1990; 1992; 1995; Bruell et al., 1992; Hamed et al., 1991; Hicks and Tondorf, 1994; Pamukcu et al., 1990; Pamukcu and Wittle, 1992; 19 93; 1993a; b; Pamukcu, 1994; Runnels and Wahli, 19 93; Shapiro and Probstein, 19 93; Ugaz et al., 1994; Wittle and Pamukcu, 19 93) However, unlike a treatment barrier, electrokinetically... 42.5 93. 9 66.9 96.0 67.4 36 .3 64 .3 86.6 89.4 93. 5 54.7 — 93. 6 — 92 .3 39.8 64.4 54.7 98.0 97.5 96.8 90.5 78 .3 95.9 83. 0 99.0 33 .0 54.5 a KS: kaolinite; KG: kaolinite and simulated groundwater; KH: kaolinite and humic substances; MS: montmorillonite; SS: clayey sand From Wittle and Pamukcu, 19 93 In natural soils with high buffering capacity and carbonate content, or those that are under the groundwater table,... 0.11 0.55 0.02 3 1 0 20 40 60 80 100 TABLE 3. 13 Some Experimental Conditions of the Electroosmosis Tests with Phenolic Compounds Test No Blank PhI PhII PhIII PhIV Contaminant Concentration (ppm) Potential Gradient Applied (V/cm) None Phenol 2-Chlorophenol 3- Chlorophenol 4-Chlorophenol 0 166 1 43 1 43 1 43 1.2 1.2 1.2 1.2 1.2 The soil was a combination of Ottawa sand (U.S Silica Company) and Georgia kaolinite... Anode End S1 = 81%; PVF = 2 .3 S2 = 53% ; PVF = 0.4 S3 = 65%; PVF = 0.1 pH FIGURE 3. 1 .3 Post electrokinetic treatment distribution of Na in drilling mud soil of various initial water saturation degrees (S1, S2, S3) (Reprinted from J Haz Mat., 55, Pamukcu, S., Weeks, A., and Wittle, J.K., Electrochemical separation and stabilization of selected inorganic species porous media, 30 5 31 8, copyright 1997, with... and Ocean Eng Convention, San Diego, Cal., Preprint 2641 Kuo, S and Baker, A.S., (1980), Sorption of copper, zinc, and cadmium by some acid soils, Soil Science Soc Am J., 44, 96 9-9 74 Lageman, R., (1989), Theory and Practice of Electro-Reclamation, NATO Pilot Study: Demonstration and Remedial Action Technologies for Contaminated Land and Ground Water, NATO/CCMS, Copenhagen, Denmark Lageman, R., 19 93, .  83. 0 Sr(II) 97.8 99.5 96.0 92 .3 99.0 U(V) 79 .3 84 .3 67.4 39 .8 33 .0 Zn(II) 54.6 43. 3 36 .3 64.4 54.5 a KS: kaolinite; KG: kaolinite and simulated groundwater; KH: kaolinite and humic sub- stances;. 64 .3 54.7 Cd(II) 94.6 98.2 92.7 86.6 98.0 Co(II) 92.2 93. 9 95.9 89.4 97.5 Cr(VI) 93. 1 94.8 97.6 93. 5 96.8 Cs(I) 71.9 80.1 74.7 54.7 90.5 Hg(II) 26.5 13. 1 42.5  78 .3 Ni(II) 88.4 95.4 93. 9 93. 6. and cathodic pH were 2.50 and 11.51, and 2.45 and 11 .34 , respectively, for samples tested with and without the GS caps. Typically during an EK test, three electrodes designated P1, P2, and P3