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a Environmental Technology Division, Pacific Northwest National Lab., P.O. Box 999, MS K9-33, Richland, WA 99352
b Dep. of Agronomy and Soils, Auburn Univ., Auburn, AL 36849
c Environmental Molecular Sciences Lab., Pacific Northwest National Lab., P.O. Box 999, Richland, WA 99352
* Corresponding author (mart.oostrom{at}pnl.gov)
Received 21 October 2005.
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ABSTRACT |
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 TOPEXECUTIVE SUMMARY ABSTRACT INTRODUCTIONAQUEOUS DISSOLUTIONENHANCED REMEDIATIONDISCUSSION AND RESEARCH...REFERENCES |
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Abbreviations: 1-D, one-dimensional • 2-D, two-dimensional • 3-D, three-dimensional • BTEX, benzene, toluene, ethylbenzene, xylene • DCA, dichloroethane • DMD, density-modified displacement • DNAPL, dense nonaqueous phase liquid • ISCO, in situ chemical oxidation • LNAPL, light nonaqueous phase liquid • NAPL, nonaqueous phase liquid • PCE, tetrachloroethene • REV, representative elementary volume • SEAR-NB, surfactant enhanced aquifer remediation at neutral bouyancy • SVE, soil vapor extraction • TCA, trichloroethane • TCE, trichloroethene
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INTRODUCTION |
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TOPEXECUTIVE SUMMARYABSTRACTINTRODUCTIONAQUEOUS DISSOLUTIONENHANCED REMEDIATIONDISCUSSION AND RESEARCH...REFERENCES |
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Bench-scale laboratory research on NAPL flow and remediation occurs at the pore scale in micromodels, the one-dimensional column scale, and the multidimensional intermediate aquifer model scale. Column studies are important because they provide an important initial understanding of a multifluid problem; however, column studies lack representiveness to field situations (Conrad et al., 2002). For instance, dissolution of entrapped NAPL in a column does not represent field conditions, because the flushing solutions are forced through the contaminated zone. Another example is the unfavorable mobilization of DNAPL following surfactant or alcohol flushing, often an apparent phenomenon in multidimensions, but masked in column studies. Two- and three-dimensional aquifer model experiments are an important intermediary between column studies and field trials.
The focus of this review is multifluid intermediate-scale flow cell experimentation. Lenhard et al. (1995) defined the conditions needed before an experiment can be classified as "intermediate-scale": (i) the experimental configuration has to allow small-scale processes to manifest themselves at a larger scale so that their relative contributions to flow and transport phenomena can be studied and quantified; (ii) the size of the experiment has to be small enough for the environment to be controlled; and (iii) the experimental-cell dimensions have to be compatible with measurement and sampling techniques.
The primary advantage of intermediate-scale experiments is that field-scale processes can be mimicked under controlled conditions (Lenhard et al., 1995). Chevalier and Petersen (1999) noted that laboratory investigations of NAPL flow in porous media are studied under the same capillary, viscous, and buoyancy forces as full-scale systems. Flow cell experiments are also conducted to provide data sets to test and verify numerical (e.g., Waddill and Parker, 1997) and analytical (Chrysikopoulos et al., 2000) multifluid flow and transport models. Waddill and Parker (1997) argued that the controlled setting and laboratory instrumentation would reduce the uncertainty in parameter estimation, thus allowing the comparisons between model simulations and experiments to focus on flow processes. Another purpose of intermediate-scale experiments is to provide the necessary data to compute parameters associated with theoretical models. For instance, parameter values for Gilland–Sherwood kinetic mass transfer models of entrapped NAPL were computed by Saba and Illangasekare (2000), and Brusseau et al. (2002) using data from flow cell experiments. Conrad et al. (2002) stressed the importance of performing flow cell visualization experiments to evaluate the performance of proposed techniques for DNAPL remediation.
An earlier literature review of 2-D (two-dimensional) laboratory experiments in NAPL flow, transport, and remediation was presented by Chevalier and Petersen (1999). They reviewed about 20 papers with the most recent contribution from 1997. Besides the fact that numerous papers have appeared in the literature after 1997, the review by Chevalier and Petersen (1999) did not provide a comprehensive review of intermediate-scale flow cell experiments before 1997.
We have reviewed about 125 original contributions. Depending on the main topic of research, the papers were divided into the following categories: (i) aqueous dissolution, (ii) enhanced remediation, (iii) flow behavior, (iv) quantification, and (v) imaging. The dissolution category includes experiments where NAPL removal occurs due to water flushing. The enhanced remediation section contains experimental contributions investigating surfactant flushing, alcohol flushing, surfactant and alcohol flushing combinations, dense brine strategies, hydraulic NAPL removal, soil vapor extraction, air sparging, heat-based remediation, bioremediation, and other techniques. The flow behavior papers involve investigations of NAPL flow and transport phenomena. The smaller quantification and imaging categories discuss experiments conducted to detect and quantify NAPL using tracers and experiments performed to provide images of fluid saturations in flow cells, respectively.
In this review, aqueous dissolution and enhanced remediation intermediate-scale experiments are discussed. In the companion paper, flow cell experiments focusing on flow behavior, quantification, and imaging are reviewed. A total of 17 original flow cell research papers dealing with NAPL dissolution are listed in Table 1. The specifics of 54 enhanced NAPL remediation papers are listed in Table 2. When experiments described in a conference paper were later published in a journal paper, only the journal paper is listed in the overview tables. For example, the NAPL dissolution journal paper by Saba and Illangasekare (2000) was included in Table 1 instead of the conference paper by Illangasekare and Saba (2000) discussing the same experiments, while the Oostrom et al. (2005a) conference paper dealing with soil vapor extraction was superseded in Table 2 by the Oostrom et al. (2005b) journal paper.
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Throughout the two reviews, we have tried to keep the terminology consistent, independent of what has been used by the various researchers. For example, we will call the discussed intermediate-scale experiments flow cell experiments, recognizing that they have been called several names, including physical aquifer model (Field et al., 1999), lab-scale reactor (Rogers and Ong, 2000), soil tank (Waduge et al., 2004), model sand aquifer (Van Stempvoort et al., 2002), sand pack (Ho and Udell, 1992), 2-D box (Taylor et al., 2004), trough (Schwille, 1967), sand box (Dai and Reitsma, 2002), soil flume (Saba and Illangasekare, 2000), and flow tank (Li and Schwartz, 2004). In addition, following the terminology proposed by Lenhard et al. (2004), we will refer to pore-scale NAPL that is occluded by water as entrapped NAPL, and pore-scale NAPL that is not entrapped by water, but does not drain from the pore spaces, as residual NAPL. Both forms of NAPL are assumed to be immobile. Residual NAPL can be continuous or discontinuous throughout the pore spaces, but entrapped NAPL is always discontinuous (Lenhard et al., 2004). The term macroentrapped NAPL is used to describe NAPL in systems where free NAPL is located in a relatively coarse porous medium surrounded by finer grained materials. Examples of such configurations were discussed by Manivannan et al. (1996), Powers et al. (1998), and Nambi and Powers (2000).
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AQUEOUS DISSOLUTION |
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TOPEXECUTIVE SUMMARYABSTRACTINTRODUCTIONAQUEOUS DISSOLUTIONENHANCED REMEDIATIONDISCUSSION AND RESEARCH...REFERENCES |
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The main goal of the experiments by Schwille (1988) was to create a defined pool of either 1,1,1 TCA (trichloroethane) or TCE on the bottom of a glass flow cell. Schwille (1988) selected 1,1,1 TCA for its intermediate solubility (720 mg L–1 at 25°C) in his first experiment, and TCE for its relatively high solubility (1100 mg L–1 at 25°C) in his second experiment. The DNAPLs, 2 L in the first and 2.5 L in the second experiment, were added directly to the bottom through two vertical tubes connected to perforated horizontal tubes. The flow cell was then uniformly filled with a sand with a saturated hydraulic conductivity of 3.5 x 10–3 m s–1. For both experiments, an even DNAPL distribution was visually observed. Water was subsequently injected and extracted through perforated tubes, positioned on either side of the flow cell, 16 cm above the bottom. No sampling ports were installed in the flow cell and only effluent samples were analyzed. For each experiment, the water injection rate was varied with time. Both experiments yielded nearly positive linear relationships between flow rate and DNAPL removal rate; however, Schwille (1988) stated that the results were not of adequate detail to allow the determination of a functional relationship governing DNAPL removal under the conditions of his experiments.
To study dissolution kinetics of TCA and TCE–TCA mixtures, Whelan et al. (1994) developed a procedure to create DNAPL pools in a glass tray, located at the bottom in a flow cell. A very useful section in their paper contains the descriptions of various failed techniques before a successful one was developed. The lower part of the tray contained a 1.6-cm-thick gravel layer, while the remainder was filled with aquifer sand. The formed pools had a distinct but not perfect flat upper surface. Dissolution experiments with TCE and TCE–TCA mixtures yielded concentrations much lower than the respective solubilities, even at sampling locations 6.2 cm above the glass tray. Dissolved concentrations at these locations were only 18% of the solubility of TCA. For binary mixtures, similar results were observed. Pool excavation showed that the fraction of sand penetrated by DNAPL at each layer increased with decreasing distance from the bottom, revealing that the researchers were not able to create a DNAPL pool with a perfectly flat surface.
Pearce et al. (1994), in a continuation of experiments discussed by Whelan et al. (1994), studied the factors that affect the dissolution dynamics of single-component DNAPL pools. Similar to what was found by Whelan et al. (1994), under conditions of steady and uniform horizontal flow, the dissolved DNAPL concentrations decreased with increasing vertical distance from the pool. The concentration gradient increased with increasing groundwater velocity. The vertical extent of the dissolved plume increased with increasing distance along the pool and decreasing groundwater velocity. In all cases, measured concentrations were a small fraction of the respective DNAPL solubility. A steady-state pool dissolution model, described by Hunt et al. (1988), for a pool with length L, was applied to compute transverse dispersivity values. The model is given by
 | [1] |
is the average pore water velocity in the longitudinal direction [L T–1], C is the dissolved NAPL concentration [M L–3], x and z are the longitudinal and vertical Cartesian coordinates [L], respectively, and Dz the transverse hydrodynamic dispersion coefficient [L2 T–1]. This coefficient can be written as Dz = 
Dm + 
zv, where is the tortuosity, Dm is the molecular diffusion coefficient, and z is the transverse dispersivity. For the boundary conditions:
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x L can be written as (Hunt et al., 1988)
 | [2] |
Chrysikopoulos et al. (2000) designed a flow cell with a well-defined circular pool. This design was in response to the pool characteristics used by Whelan et al. (1994) and Pearce et al. (1994). Chrysikopoulos et al. (2000) commented, "The experimental protocols developed by Whelan et al. (1994) do not lead to a well-defined and flat pool–water interface. Formation of entrapped DNAPL ganglia in the upper half of the glass pan may occur during injection. Accurate determination of the pool–water interfacial area resulting from this setup is not a trivial task. Furthermore, the intrusion of a glass pipet in the aquifer may disturb the interstitial flow." The new pool design consisted of a bottom plate with a 0.5-cm-deep and 7.6-cm-diameter circular depression, which was filled with 0.5-cm-diameter glass beads. The TCE was then added to the depression through a tube connected to a reservoir to form a pool. The void volume of the pool was estimated to be 
12 cm3. Throughout the flushing experiments, the pool was kept filled with TCE. Hence, pool depletion was not considered. Another drawback of this design is that the TCE was not in the porous medium itself. The flow cell was packed with a kiln-dried Monterey sand under a water head of 5 cm. Liquid samples were taken during steady-state conditions from seven ports placed at different depths, primarily on the center axis of the flow cell. A total of three experiments were completed with different pore water velocities. The steady-state concentration distributions were analyzed using the analytical solution by Chrysikopoulos (1995) for a transport equation including retardation, dispersion, and kinetic dissolution. All transport parameters for this model were obtained independently, except for the average kinetic mass transfer coefficient. This coefficient could not be determined independently but was fitted using concentrations from a subset of the ports. The concentrations at the remaining ports, all located on the centerline of the aquifer, were then predicted using the analytical solution. The analytical solution was found to describe observed TCE plumes well. Important conclusions from the experiments were that observed aqueous TCE concentrations just 15 to 60 cm downgradient from the pool, and just a few centimeters above the upper boundary of the pool, were <4% of the aqueous solubility. According to the researchers, this observation provides strong evidence in support of the notion that rate-limited dissolution of DNAPLs can be the cause of contaminant plumes that are dominated by concentrations several orders of magnitude less than the aqueous solubility. In addition, it was stated that the results also imply that observed concentrations exclusively at the parts per billion level do not preclude the existence of a DNAPL pool at a site.
The experiments reported by Dela Barre et al. (2002) were conducted in the same flow cell used by Chrysikopoulos et al. (2000); however, Dela Barre et al. (2002), introduced the DNAPL PCE (tetrachloroethene) in the porous medium atop impermeable (glass) and relatively impermeable (clay) surfaces at the bottom plate of the cell. The objectives of this study were to observe the controlled dissolution of DNAPL pools into an overlying aquifer, and to rigorously test an existing mathematical model (Chrysikopoulos, 1995) describing the dissolution of an ideally configured pool. In the first experiment, a pool containing 1 cm3 of PCE was placed on the bottom of the flow cell. In the second experiment, the pool was placed within a depression in the 1-cm clay layer that was placed on the bottom of the cell. The pools were irregular in shape, finite in mass, and resided within the porous medium, where interfacial forces were allowed to affect the pool surface. The porous medium was the same sand used by Chrysikopoulos et al. (2000). In the experiments, temporal sampling was conducted at a designated observation point. After steady-state conditions were reached, samples were obtained from up to 35 sampling locations. For the first experiment, variable flow rates were used, while for the second experiment, the flow rate was kept constant. As in the modeling approach by Chrysikopoulos et al. (2000), most of the parameters for the model were obtained independently, except for the value of the mass transfer coefficient. The researchers discovered that, despite the finite pool volumes, observed plumes achieved a quasi-steady-state distribution during extended time periods and that the effect of mass loss on pool geometry was minimal during the course of the experiments. It was also found that pool dissolution could be modeled reasonably well with a boundary layer theory assuming a simplistic circular pool and adjustment of a pool-average mass transfer coefficient. Modeling discrepancies were the largest closest to the pool due to the irregularities of the pool. An interesting observation was that the estimated mass transfer coefficient values were three to four times greater than those estimated in the dissolution study by Chrysikopoulos et al. (2000), involving an ideally configured pool, and two to three times greater than values predicted by a theoretically based mass transfer correlation for elliptical or circular pools. It was suggested that the most likely cause of these discrepancies was the pool dissolution model's failure to address interfacial issues associated with the emplaced pool and overcompensation in the form of elevated mass transfer coefficients. The researchers stated that, partly based on findings from pore-scale research, pore-scale features of a DNAPL pool–water interface will impact dissolution and may change significantly as sufficient portions of the pool dissolve. They concluded that future research is needed focusing on the connection between pool dissolution and pore-scale mechanisms defining the pool–water interface.
The only LNAPL pool dissolution study was conducted by Voudrias and Yeh (1994). A rectangular pool, containing 1 L of dyed toluene, was constructed directly on top of the water table. One horizontal row of sampling ports was located 3 cm below the pool–water table interface. The experiment consisted of six stages with variable pore water velocity. The duration of each stage lasted from 48 to 500 h. The model of Hunt et al. (1988) was applied to the dissolved plumes under steady-state conditions. Transverse dispersivities ranged from 0.013 to 0.018 cm, which are comparable to the values computed by Pearce et al. (1994). The researchers also observed very steep vertical concentration gradients in groundwater flowing horizontally below the dissolving toluene pool. It was estimated that, under the current experimental conditions, the removal time would be between 8 and 12 yr.
An intermediate-scale flow cell experiment to study the flow of liquid and the transport of dissolved TCE (trichloroethylene) in a saturated, heterogeneous porous medium system was conducted by Oostrom et al. (1999a). In contrast to the other contributions discussed in this section, the pool was not directly emplaced but resulted from a spill from the top of the flow cell. The flow cell was packed with three layers and five lenses consisting of four different sands. All lenses and layers had horizontal interfaces, except the lowest interface, which was pointed down in the middle. Horizontal groundwater flow was imposed by manipulating the water levels in two end chambers. More than 0.5 L of dyed TCE was allowed to infiltrate at a constant rate into the porous medium from a narrow source located on the surface. Fluid samples were collected from 20 sampling ports to determine dissolved TCE concentrations. Visual observations and measured TCE saturations indicated that the spilled TCE accumulated on top of, but did not penetrate into, fine-grained sand lenses and layers but that some TCE infiltrated into medium-grained sand lenses. Most of the TCE finally pooled on top of a fine-grained sand layer located near the bottom of the flow cell. The simple Hunt et al. (1988) pool dissolution model was used to predict observed dissolved TCE concentrations. Results showed that the measured concentrations could only be predicted with unrealistically high transverse dispersivity values. The main reason for the discrepancy was that the observed TCE concentrations were the result of contributions of entrapped and pool dissolution.
The only dissolution study with a field DNAPL pool was conducted by Eberhardt and Grathwohl (2002). The study presented results in the largest intermediate-scale flow cell (7.43 m long by 2.75 m wide by 1.0 m high) reported here. The flow cell contained both a 2.5-m-long by 0.05-m-wide by 1.0-m-high coal tar pool and a 0.5-m-long by 1.0-m-high by 1.0-m-wide zone with entrapped NAPL. The entrapped NAPL dissolution is discussed below. The coal tar contained contaminants like BTEX (benzene, toluene, ethyl-benzene, xylene) and PAHs (polycyclic aromatic hydrocarbons). Three aquifer materials were used. A low-permeability sand was used at the bottom and below the entrapped zone. A medium-grained sand and a fine gravel were used to build up the main aquifer and a high-conductivity zone, respectively. The DNAPL pool was made by first excavating a 3-cm-deep depression in the fine sand layer at the bottom. The depression was then filled with aquifer sand, followed by emplacement of 40 kg (33.3 L) of coal tar. The coal tar pool had a 100% NAPL saturation. The total number of sampling ports for this large cell was only 46, which is remarkably low for a flow cell of this size. The flow velocity in the cell was 1.7 m d–1 for the first 177 d. After that, it was increased to 5.1 m d–1 for 42 d, then reduced to 3.4 m d–1 for 21 d, and finally further reduced to 1.7 m d–1 for the final 114 d. The total duration of the experiment was 354 d. The objectives of the pool dissolution study were to determine the validity of Raoult's law for prediction of aqueous concentrations and to quantify the contribution of the transverse vertical dispersion on BTEX and PAH dissolution from a pool. Results showed that Raoult's law was applicable for estimation of aqueous concentrations, assuming an activity coefficient of 1. Calculations based on a steady-state pool dissolution model (Hunt et al., 1988) showed that transverse vertical mixing, with a dispersivity value of 0.01 cm, dominated the pool dissolution process.
Dissolution of Entrapped Nonaqueous Phase Liquids
Anderson et al. (1992) and Reynolds et al. (1996) investigated dissolution of entrapped, circular source zones. The work of Imhoff et al. (1996) focused on dissolution fingering. Saba and Illangasekare (2000) used similarly shaped LNAPL zones to investigate rate-limited dissolution. The flow cell experiments reported by Brusseau et al. (2000, 2002) were designed to study the rate-limited dissolution of rectangular source zones with entrapped DNAPL. The only flow cell dissolution experiment with entrapped field NAPL was discussed by Eberhardt and Grathwohl (2002).
The emphasis of the research by Anderson et al. (1992) was on obtaining mass removal rates when groundwater is free to flow at least partially around a zone of a porous medium that contains entrapped solvent. The experiments were conducted in response to results from 1-D (one-dimensional) laboratory column studies, which indicated that when water is forced to flow through a zone with ample entrapped DNAPL, saturation concentrations can be achieved rapidly. The zone with entrapped DNAPL (PCE) had a cylindrical shape with a diameter of 15.2 cm and was placed in the middle of the flow cell. First, a sheet-metal cylinder was placed vertically into the center of the flow cell. The flow cell and cylinder were then filled with sand. The water table was set at 74 cm above the bottom. To produce the entrapped zone, dyed PCE was injected into the cylinder from the bottom until the PCE was even with the water table. Mobile PCE was subsequently siphoned off and water was again injected from the bottom to displace the remaining mobile PCE. Finally, the sheet metal was removed from the flow cell. This procedure yielded a zone with 13% entrapped PCE. A total of 81 sampling ports were installed in three horizontal rows and one vertical column at the effluent end of the tank. Experiments were completed at pore water velocities of 10, 30, 60, and 100 cm d–1. Observed steady-state concentrations for all experiments were nearly similar. PCE concentrations in the center of the dissolved plume, emanating from the source zone, were at or near saturation, even at water velocities of 100 cm d–1. A main conclusion of the experiment was that reduction of mass removal due to reduced permeability in the solvent zone was found to be minimal.
The main objective of the experimental work by Reynolds et al. (1996) was to assess if mass transfer coefficients calibrated from 1-D dissolution data can be used to accurately model 3-D flow and dissolution. Circular sources containing entrapped toluene were used. It was observed that, contrary to the other work reported here, concentrations increased with distance. The reasons that were considered for this unexpected behavior were channeling within the sand, heterogeneous dissolution within the zone, and global flow deviation. Excavations of the cell did not provide support for the channeling suggestion. The second reason, isolated ganglia being dissolved and forming "mini-plumes" was thought to be the most likely. A modeling effort was able to duplicate the general trends exhibited by the experimental data, although adjustments were needed of the transverse dispersivity and relative permeability data.
Imhoff et al. (1996) used a relatively small 2-D flow cell to study the effects of porous medium structure, Darcy flux, initial entrapped NAPL saturation, median particle diameter, gravity, and NAPL composition on dissolution fingering through preferential flow paths. The model for dissolution fingering of a single component NAPL presented by Imhoff and Miller (1996) indicated that fingering depends only on the pore water velocity and the entrapped NAPL saturation; however, the remaining parameters were investigated for the following reasons: (i) porous medium structure was hypothesized to affect finger location; (ii) particle size would affect the size of ganglia and the length of the dissolution front, which might inhibit finger formation; (iii) gravity was assumed unimportant in the model; and (iv) NAPLs in the environment are frequently mixtures and it was important to show that dissolution fingering also occurs for mixed systems. Experimental observations showed that fingering occurred when two conditions were met: (i) 11 to 80 e-fold times had elapsed, where e-fold time is the time required for the instability to grow by a factor e and was predicted by the linear stability analysis outlined in a theoretical companion paper (Imhoff and Miller, 1996); and (ii) the length of the dissolution front before finger development was smaller than the zone with entrapped NAPL. The latter condition is not part of the theoretical model, because it assumes a thin dissolution front. It was further observed that Darcy flux and median particle diameter influenced the length of the dissolution front. In addition, results showed that fingers grew faster for smaller initial entrapped NAPL saturations and larger Darcy fluxes, were not affected by gravity, and occurred in an experiment with a TCE–toluene mixture. Variation of the finger growth rate with initial degree of NAPL saturation and Darcy flux are in agreement with the analysis of Imhoff and Miller (1996). In a 3-D experiment, fingers were found to grow >30 cm in length, while theoretical predictions indicated that they might grow to meters in length, depending on the flow rate, transverse mixing, and finger radius. Predictions using existing 1-D correlations not including fingering were shown to be significantly different from experimental observations of TCE dissolution. These results suggest that dissolution fingers can dramatically affect the distribution of NAPL saturations and aqueous phase solute concentrations; however, Imhoff et al. (1996) noted that when cleanup was defined as the disappearance of visible TCE ganglia, dissolution fingers led to a cleanup time that was only 10% longer than the predicted time based on the local equilibrium assumption. At the scale of these experiments, the effect on cleanup time is not large, but further study is recommended to assess the importance of dissolution fingering on cleanup time at the field scale.
The flow cell work by Saba and Illangasekare (2000) was motivated by the observation that mass transfer from entrapped NAPL in the subsurface takes place in 3-D groundwater flow fields but that most dissolution laboratory studies had been only 1-D. They argued that multidimensional studies were needed to capture more realistic conditions such as the heterogeneous distribution of entrapped NAPL ganglia and the potential for flow bypassing due to reduced water permeability. The main objective of this study was to evaluate the effects of flow dimensionality on NAPL dissolution in experiments using 2-D flow fields. Sources containing entrapped para-xylene were placed in an otherwise homogeneous sand pack. The packing procedure of the source zones is of interest. A 20-cm-long, 5.08-cm-high, and 5.08-cm-wide aluminum mold was filled with water-saturated sand. The mold end plates were coated with stainless steel screens and the internal walls were coated with greased gaskets to avoid wall effects. Para-xylene was injected into the mold until free phase LNAPL appeared in effluent. The saturation was found to be 0.85. To create a zone with entrapped NAPL, two pore volumes of deaired water were subsequently pumped into the mold at a low rate until the LNAPL was no longer displaced, as verified by visual observation of the effluent. The volume-averaged entrapped saturation was calculated to be 0.22. Uniform distribution throughout the sand was verified by dividing a test mold into six sections and obtaining the LNAPL saturation of each piece using ultraviolet spectroscopy. The results show that the saturations were reasonably constant for all sections. After the entrapped LNAPL was formed, the molds were stored in a freezer until they were emplaced in the flow cell. Following a tracer test to determine the dispersivity of the 0.60-mm sand, dissolution behavior under different groundwater velocities and entrapped NAPL source lengths (5.08, 10.15, and 15.24 cm) were investigated. The groundwater velocities ranged from 0.25 to 3.0 m d–1 for each source length. A column of sampling ports was located a short distance downstream from the entrapped NAPL zone. In the dissolution experiments, the flowing aqueous phase appeared to infiltrate through the contaminated zones as multiple fingers rather than a uniform front. It was expected that concentrations would decrease with increasing velocity but the opposite was observed. This behavior was attributed to fingering of the aqueous phase through the contaminated zone, a concept introduced by Imhoff et al. (1996). A phenomenological model development yielded a relation that is expected to govern the mass transfer inside a REV (representative elementary volume; Saba and Illangasekare, 2000):
 | [3] |
/
, and Sc is the Schmidt number given by Sc = /Dm. The mass transfer coefficient kln [T–1] is a lumped factor given by kln = Klnaln where Kln is the average mass transfer coefficient [L T–1] for the aqueous phase (subscript l) and NAPL (subscript n) interface, and aln is the specific interfacial area [L–1] between entrapped NAPL and groundwater. Furthermore, a, ß, , and 
are fitting parameters, dp is the mean particle diameter [L], Dm is the molecular diffusion coefficient [L2 T–1], is the aqueous phase velocity [L T–1], is the tortuosity of the sand, 
n is the volumetric NAPL content, L* [L] is the path length the aqueous phase travels in the source zone, and is the kinematic viscosity [L2 T–1] of water. The mass transfer coefficient is found in a commonly used form of the mass transfer relationship:
 | [4] |
nd50/L* was proposed to represent the changing geometry in a contaminated zone block when NAPL is dissolved with time. The definition of L* is not completely clear but Saba and Illangasekare (2000) stated that in numerical simulations, L* is the length of the contaminated numerical soil block in the flow direction. This phenomenological model was implemented in the transport code MT3D (Zheng, 1990) and the dissolution experiments were simulated using a combination of MODFLOW (McDonald and Harbaugh, 1988) and MT3D. The values of all the parameters, except for the fitting parameters a, ß, and , were obtained independently. The latter were obtained using the inverse model UCODE (Poeter and Hill, 1998). Saba and Illangasekare (2000) found that dissolution could be simulated using the relationship
 | [5] |
Brusseau et al. (2000) noted that aqueous phase concentration data often serve as the basis of decisions regarding remediation but that aqueous concentrations of NAPL constituents are often orders of magnitude lower than their equilibrium concentrations, even at sites where their presence has been confirmed. This severely constrains the use of concentration data for determining mass distributions. Knowledge of NAPL dissolution behavior under conditions prevalent at the field scale is critical for conducting risk assessments and implementing appropriate remediation strategies for NAPL-contaminated sites. In addition to pore-scale rate-limited mass transfer, factors that cause apparent rate-limited behavior are the result of limited contact between advecting water and NAPL. For instance, when bypass flow occurs due to reduced permeability, NAPL dissolution is affected. This situation might occur where entrapped NAPL resides in low permeability zones or when pooled NAPL severely reduces the aqueous permeability. Dilution in wells might also significantly reduce observed concentrations. The result of research shows that the magnitude of aqueous concentrations in the presence of NAPL can be significantly influenced by NAPL distribution, porous-media heterogeneity, and sampling method.
Dissolution experiments were conducted using an intermediate-scale flow cell packed with a medium-grained sand in which two zones of entrapped TCE saturation were emplaced (Brusseau et al., 2000). One (Zone 2) was created in the same medium-grained sand as used for the flow cell matrix, and the other was created in finer sand (Zone 1). Aqueous samples were obtained using depth-specific sampling, vertically integrated sampling, and the extraction well. A dual-energy radiation system was used to determine entrapped DNAPL saturations before and during the dissolution process. A tracer test (Br) was done to examine the potential influence of nonuniform flow and dilution-related factors on solute transport. The tracer solution contained a fluorescent dye for visual observations. Results show that nonuniform flow behavior was associated with the TCE zones. Specifically, the front was delayed slightly by Zone 2 and significantly by Zone 1. For Zone 2, dissolution occurred relatively uniformly across the upgradient edge and continued progressively along the longitudinal axis of the zone. For Zone 1, mass removal was limited to the periphery. For this zone, the relative permeability was 80 times smaller. The NAPL dissolution and mass removal were strongly affected by heterogeneity. The results of this study indicate that nonuniform NAPL distribution, physical heterogeneity, and associated nonuniform flow considerably influenced the magnitude and location of NAPL dissolution. The magnitude of aqueous phase concentrations varied as a function of location and sampling method. The concentrations at the vertically integrated ports and the extraction well were less than the concentrations measured at the point-sampling ports, indicating dilution effects. This illustrates the importance of considering the type of sampling method used when analyzing and interpreting data. The smaller concentrations were caused by sampling-associated dilution rather than by rate-limited interphase mass transfer at the pore scale. The results reinforce the concept that the presence of a NAPL cannot be necessary ruled out even when aqueous concentrations are only a fraction of expected equilibrium concentrations (Brusseau et al., 2000).
Brusseau et al. (2002) described an experiment similar in design to the one discussed by Brusseau et al. (2000). In the second experiment, the lower zone (Zone 1) consisted of a 0.85 to 0.60-mm sand with entrapped TCE. The upper zone (Zone 2) contained entrapped DCA (dichloroethane). The solubility of DCA is about 8.6 g L–1 at room temperature vs. 1.1 g L–1 for TCE. A 3-D mathematical model was used to solve the governing equation for solute transport with dissolution of an immiscible liquid, described using the widely used first-order mass transfer equation. The Sherwood correlation used by Brusseau et al. (2002) is the one proposed by Powers et al. (1994):
 | [6] |
= d50/dM is a normalized grain size with dM = 0.05 cm (reference diameter), Ui = d50/d10 is the uniformity index for the porous medium, d50 is the diameter of the media grains, 50% of which in weight are smaller than d50, and ß are coefficients, and n0 is the initial NAPL saturation, as determined with the dual-energy radiation system. Other parameters and variables in Eq. [6] were defined above. Column experiments were completed to independently determine initial dissolution rate coefficients. The columns were packed and flushed using similar methods as were used for the entrapped source zones in the flow cell. A 1-D mathematical transport model coupled to an optimization program was used to obtain initial dissolution rate coefficients. With the initial rate coefficients, the value in Eq. [6] could be computed. Brusseau et al. (2002) used the Sherwood equation differently than Saba and Illangasekare (2000). Equation [6] was not used to solve the inverse problem to obtain kln values. Instead, Eq. [6] was used to convert the initial values obtained from the column experiments to discrete nodal values, accounting for spatial differences in the physical properties. Specifically, the initial mass transfer coefficient values for the flow cell experiment were computed as
 | [7] |
 | [8] |
The only dissolution study of entrapped field DNAPL was conducted by Eberhardt and Grathwohl (2002). Details of this experiment were discussed above. The entrapped zone was constructed by mixing known amounts of DNAPL and wet porous media. A total of 11.4 kg (9.5 L) of coal tar was used to create the entrapped zone, resulting in a saturation of 4.96%. Regarding the entrapped DNAPL dissolution, the objectives of the study were to determine the validity of Raoult's law for prediction of aqueous concentrations, and to test the local equilibrium assumption for entrapped coal tar dissolution. Results showed that Raoult's law was applicable for estimation of aqueous concentrations, assuming an activity coefficient of 1. In addition, it could be demonstrated that the local equilibrium assumption could be used for an entrapped zone of this size. No significant decrease of the fluxes of any of the compounds was observed during the total run time of the experiment, indicating that no additional mass transfer resistances arose. Some dissolution fingering was observed during the entrapped NAPL dissolution.
Dissolution of Macroentrapped Nonaqueous Phase Liquids
The work by Manivannan et al. (1996), Powers et al. (1998), and Nambi and Powers (2000) addressed processes governing the overall dissolution of the various NAPLs entrapped in a coarse sand lens in an otherwise fine-grained sand at a high saturation. Mannivannan et al. (1996) used a coarse-grained rectangular lens with dimensions of 3 by 2 by 2 cm. The entrapment in the coarse sand lens was established through immiscible displacements using the entire pore volume of the cell. This procedure yielded an average saturation of 0.10 to 0.20 in the fine-grained and 0.80 in the coarse-grained lens. After the formation of the two-phase system, the flow cell was flushed with clean water. A series of photographs was presented to document the dissolution process. First, it was observed that TCE was selectively removed around the periphery of the lens, indicating a bypassing of the lens due to a reduced aqueous phase permeability. Later, the dissolution also started to occur in the lens. Apparently, in addition to water bypassing around the DNAPL-contaminated coarse sand lens, sufficient water flowed through this region to cause an increase in relative permeability and effluent concentration with time. Two simple mathematical models were developed to explore possible physical phenomena contributing to the observed phenomena. The assumptions for the first model were that the DNAPL in finer sands dissolves first, followed by some flow of water through the coarse lens, leading to, finally, a greatly reduced DNAPL saturation in the lens. The second model assumes that water–DNAPL contact only occurs along the perimeter of the lens and that, as the DNAPL saturation in the contaminated lens is reduced, a larger fraction of the water will flow through the lens. Manivannan et al. (1996) showed that the first model produced the better results.
The experiments described by Powers et al. (1998) are a continuation of the work by Manivannan et al. (1996). For the 2-D studies, TCE, DCA, and o-toluidine were used in separate experiments. The nearly neutral density DNAPL o-toluidine, with a solubility of 16 500 ppm, was introduced to further isolate the dissolution processes from the multiphase flow phenomena caused by gravitational forces. The coarse sand lens was larger (5 by 2 by 3 cm) than in the study by Manivannan et al. (1996), and a different fine-grained sand was used. The emplacement process was changed to direct injection of a known volume of NAPL through a port installed in the back of the cell. This procedure created a high degree of NAPL saturation in the lens, while the surrounding fine-grained sand stayed clean. The 2-D experiments were designed to isolate processes controlling dissolution of lenses containing high saturations of NAPL. The experimental variables were aqueous phase velocity, type of NAPL, grain size of the coarse lens, and initial NAPL saturation. For a typical experiment with o-toluidine, photographic evidence showed that as time progressed, a reduction occurred in the overall NAPL saturation in the lens without an apparent spatial variability in saturation; however, random heterogeneities were apparent. More variability was observed using the denser DNAPLs TCE and DCA as a result of severe gravity drainage. For this reason, the experiments with TCE and DCA were not continued. Quantitatively, the effluent curves showed a gradual increase in concentration for the first 60 to 80 pore volumes, followed by a steep decline. The modeling combined MODFLOW (McDonald and Harbaugh, 1988) for flow and MT3D (Zheng, 1990) for transport, and was more extensive than for the Manivannan et al. (1996) study. The models were coupled by a mass balance subroutine to estimate the mass NAPL loss at each time step. The coarse sand was treated as a single REV with the NAPL saturation averaged throughout this volume. Relative permeability and dispersivities were not directly measured and the local equilibrium assumption was assumed for NAPL dissolution. As a result, Powers et al. (1998) stated that the modeling was only intended to show general trends, and not to match the data explicitly. In general, the model predictions were consistent with the overall trends in the experimental data, meaning that the simulations predict an increase in effluent concentrations, followed by a rapid decline. The model was most sensitive to estimates of the relative permeability in the coarse lens. As expected, the experimental data show more tailing than the model, which employed local equilibrium dissolution. For NAPL saturations larger than 0.15, system hydrodynamics were indicated as the sole rate-limiting process. Below this value, rate-limited mass transfer reduced the observed concentrations.
Nambi and Powers (2000) completed several additional experiments using a similar design as described by Powers et al. (1998) to understand and identify the parameters controlling NAPL dissolution in heterogeneous systems. O-toluidine was again used to ensure uniform distribution. Several parameters that influence the hydrodynamics of the system, i.e., sand grain size, size and distribution of the different permeability zones, and the initial NAPL saturation, were varied. The experiments were motivated by speculations of Powers et al. (1998) that aqueous flow patterns within a heterogeneous system are largely responsible for variable effluent concentrations in the 2-D system of concern. Initial NAPL saturation and the intrinsic permeability ratio, two variables that affect the ratio of effective permeability between the NAPL-contaminated coarse sand and the surrounding fine sand, have the most significant impact on effluent concentrations. As NAPL saturations within the coarse lens were reduced below 0.3, thermodynamic equilibrium was not achieved, resulting in a rapid decline in measured concentrations at the effluent ports. During that stage, decreases in the NAPL interfacial area and increases in the aqueous phase velocity could both contribute to mass transfer rate limitations. Another critical parameter is the exterior surface area of the NAPL-contaminated zone. A wider zone and distribution throughout two distinct zones resulted in larger aqueous phase concentrations. Mass transfer correlations for these experiments were developed by Nambi and Powers (2003).
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ENHANCED REMEDIATION |
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TOPEXECUTIVE SUMMARYABSTRACTINTRODUCTIONAQUEOUS DISSOLUTIONENHANCED REMEDIATIONDISCUSSION AND RESEARCH...REFERENCES |
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In general, surfactants are used to increase the total aqueous solubility of a NAPL or to decrease NAPL–water interfacial tension to promote mobilization (Lowe, 1997). Depending on the desired effect, appropriate surfactants need to be chosen. The chosen surfactants are delivered to entrapped and pooled regions through the aqueous solution. Surfactant molecules have both a hydrophilic and a hydrophobic site, and are classified as either anionic, cationic, or nonionic. Anionic and nonionic surfactants tend to be good solubilizers and are relatively nontoxic. Nonionic surfactants are less sensitive to the presence of salts than anionic surfactants. They are also often used as cosurfactants. Cationic surfactants tend to be toxic. At the critical micelle concentration, contaminants can partition into the interior (hydrophobic) part of the micelles, referred to as micellar solubilization (Abriola et al., 1995). Large contaminant concentration increases into aqueous solutions can thus be accomplished, which is the basis of using surfactant flushing as a promising remediation technique. The degree of solubility enhancement, characterized by the molar solubility ratio (i.e., the moles of contaminant solubilized per mole of surfactant), depends on the combination of NAPL and surfactant. These types of systems are typically referred to as Winsor Type I systems or single-phase microemulsions with the micelles residing in the water. At ultra-low interfacial tensions, a middle-phase microemulsion or Winsor Type III system can be formed. It exists as a separate phase with a density between that of water and the NAPL. Winsor Type II systems occur when the NAPL is rich in surfactant micelles containing water. The hydrophobic ends of the surfactant molecules are now pointed outward toward the NAPL. Winsor Type II systems must be avoided because of the partitioning of the surfactant into the NAPL phase. Prior testing in the laboratory is needed to obtain the optimum surfactant for a given NAPL.
The degree of interfacial tension reduction, and hence capillary force, required to mobilize NAPL can be characterized by the capillary number, the bond number, and the trapping number (Pennell et al., 1996). The capillary number, NCa, represents the ratio of viscous (mobilizing) forces to capillary (resisting) forces and is defined as
 | [9] |
nw and are the interfacial tension and contact angle, respectively, between the NAPL and the aqueous phase. The bond number, NB, represents the ratio of buoyancy (gravity) to capillary forces and is defined as
 | [10] |
is the difference between the densities of the aqueous phase and NAPL, g is the gravitational field strength, k is the permeability, krw is the relative permeability to the aqueous phase, and the other variables have been defined above. For a vertically oriented pore, with NAPL displacement in the direction of the buoyancy force, the sum of the bond and capillary numbers, referred to as the total trapping number (NT) controls the displacement (Pennell et al., 1996).
Light Nonaqueous Phase Liquid Surfactant Flushing
The first intermediate-scale experiment in which a surfactant mobilized a NAPL was the qualitative test reported by Schwille (1967). A relatively large amount of a heating oil had infiltrated directly above the capillary fringe of a highly permeable porous medium. After redistribution of the oil body, oil recovery was attempted by spraying the soil with an aqueous detergent solution. After the spraying, the oil reportedly moved quickly to a recovery ditch. Chu et al. (1996) investigated the removal of organic liquid (dodecane) contaminants with surfactant foams (Bioterge AS-40). It appeared that the foam bubbles enhanced the removal of dodecane primarily by (i) plugging clean areas through its stability in zones free of dodecane and its instability in contact with dodecane, and (ii) breaking dodecane into stable globules whose diameters are smaller than those of the pore necks through which they must travel to be removed. It should be pointed out that this experiment, although unique in nature and promising as a remediation strategy, was performed under idealized conditions in a 2-D flow cell of only 2.5-mm thickness. The results of the experiments were only discussed in qualitative terms.
Chevalier et al. (1998) tried to mobilize a gasoline spill in a variably saturated 2-D flow cell with the use of the surfactant solution dodecyl benzene sulfonate. The gasoline was positioned on top of the capillary fringe. The reduction in interfacial and surface tensions resulted in the desaturation of the aqueous phase and the subsequent flow of gasoline into previously water-saturated pores. With time, the spill transformed from a semicircle to a long, thin lens. The researchers performed an analysis based on the capillary, the bond, and the total trapping number to see if pretreatment with a surfactant could be beneficial before activating LNAPL pumping. The analysis showed that treatment with a surfactant before oil pumping would remove more LNAPL than the usual sequence of pumping followed by a surfactant treatment. Surfactant pretreatment transforms entrapped LNAPL, which cannot be removed by pumping, into free and mobile LNAPL that can be removed.
In a 2-D entrapped LNAPL zone experiment, Saba et al. (2001) showed the complete removal of entrapped p-xylene from a rectangular lens with polyoxyethylene sorbitan monooleate (Tween 80). A total of three experiments were conducted with different surfactant flow velocities and NAPL source zone lengths of 5.1 and 20.3 cm, respectively. The preparation and emplacement of the entrapped NAPL zones were identical to those of Saba and Illangasekare (2000) for aqueous phase dissolution experiments. The researchers emphasized that the contaminant dissolution occurred from the boundaries of the source toward its center. This was in apparent contrast to the unidirectional dissolution front characteristic of 1-D experiments. The primary goal of this research was to obtain additional insight into the mechanisms of mass transfer under 2-D flow fields. A total of four Gilland–Sherwood relationships were proposed to describe the mass transfer. After a theoretical analysis, one of the four models was deemed to be inappropriate because of its limited application to high NAPL saturations. The values of the parameters in the relationships of the remaining three models were determined using a method similar to the one used in Saba and Illangasekare (2000). These three models produced the same level of confidence. Unfortunately, the models were not applied to an independent experiment.
Dense Nonaqueous Phase Liquid Surfactant Flushing
Walker et al. (1998b) performed a surfactant-enhanced remediation experiment in an unsaturated coarse sand porous medium with a fine sand layer embedded. They attempted to clean up a PCE spill that had entered the porous medium with the water table near the bottom of the flow cell (Hofstee et al., 1998b). The location of the water table caused the coarse sand just below the fine sand layer to be unsaturated, while the bottom part of the fine sand layer was at or near saturation. The PCE spill entered and accumulated mostly in and below the fine sand layer. To hydraulically control the flow of the surfactant (Triton X-100) plume, the number of extraction ports and the extraction and influent pumping rates were varied. During the injection of the surfactant solution, Walker et al. (1998b) observed spontaneous spreading of the surfactant molecules. As a result of the decreased interfacial tension, the height of the saturated zone in the fine sand layer was reduced rapidly from approximately 20 to 4 cm. The researchers were able to show, through differential scanning calorimetry analysis, that the surfactant molecules prefer to reside at fluid–fluid interfaces rather than at fluid–solid interfaces. Walker et al. (1998b) also noted that drainage may be useful for air sparging or vapor extraction because more of the contaminated zone would be exposed to air. Upon entering the coarse sand below the fine sand layer, the surfactant plume formed definite fingers, which were attributed to a change in the porous medium properties at the fine–coarse sand interface and the increased density of the surfactant solution due to the solubilization of PCE. Walker et al. (1998b) demonstrated the importance of extraction well locations to promote removal of all PCE by the surfactant flushing process. They also reported the existence of milky Winsor Type I microemulsions.
Using a much more stratified porous medium and TCE as the DNAPL, Oostrom et al. (1999b) supported the basic findings of Walker et al. (1998b), i.e., unstable fingering and sinking of solutions containing solubilized TCE, formation of small TCE droplets, and milky, emulsion-like appearances. Oostrom et al. (1999b) used alternately pump-and-treat and surfactant (T-MAZ-80) flushing with a two-well system. Probably due to the added heterogeneity, only 60% of the TCE was removed. During excavation of the sand from the flow tank, free TCE was found in the bottom layer of fine sand, where it had been out of reach of the flushing solutions.
The effect of rate-limited micellar solubilization and subsurface layering on the recovery of PCE by 4% Tween 80 was clearly demonstrated by Taylor et al. (2001). They not only showed that the rate of PCE solubilization depended on the Darcy velocity, but that even under no-flow conditions, instantaneous equilibrium did not exist. Taylor et al. (2001) found that, similar to Walker et al. (1998b), the presence of fine layers resulted in an initial bimodal PCE distribution that consisted of high saturation zones or pools above fine layers and regions of entrapped PCE existing as entrapped droplets or ganglia. The entrapped NAPL was easily removed by Tween 80, but Taylor et al. (2001) encountered the most difficulty in removing the pooled PCE collected on top of the fine layers. They did not observe the entry of PCE into these fine layers. The PCE recovery was strongly affected by the low-permeability lenses as well as mass transfer limitations. The experiments showed clearly that the injected surfactant solutions, which were slightly denser than the ambient solution, flowed preferentially along the bottom of the flow cell. They argued that, for DNAPL pools above a confining layer, the use of denser surfactant solutions may be advantageous due to the tendency of such solutions to flow along the bottom of a contaminated aquifer. The experiments were simulated with the MISER code by Rathfelder et al. (2001). Model parameter values were derived from independent experiments. The model assumes an immobile PCE phase and rate-limited mass transfer for solubilization and surfactant sorption. An explicit formulation of the PCE–water interfacial area was incorporated to obtain an accurate prediction of the solubilization process. Since PCE saturations were not directly measured, the initial saturation distribution needed for the MISER simulations was estimated from visual observation or obtained from a multifluid flow simulation of the PCE injection (Rathfelder et al., 2001). The uncertainties in the initial distribution caused discrepancies between model predictions and measured values.
When a porous medium has been exposed to a NAPL for a long time, the solid phase may become hydrophobic rather than hydrophilic. Such is the case with, for example, coal tars. Dong et al. (2004) were able to show, in very simple qualitative experiments, that coal tar could be mobilized at intermediate concentrations using polymeric surfactants. If the concentration of the surfactant was too high, dissolution of the coal tar would occur. If it was too low, no effect was measurable.
The flow cell work by Ramsburg and Pennell (2001) was done to test two surfactant formulations with either Tween 80 or Aerosol MA-80 in support of source zone remediation of PCE at a former dry cleaning facility. The saturated flow cell was packed with a lower permeability zone at the bottom and two narrow zones with low-permeability lenses in an otherwise medium-grained aquifer sand. The interfacial tension of Tween 80 solution with PCE was 4.90 dynes cm–1 and the equilibrium solubility was 26 060 mg L–1. The Aerosol MA-80 interfacial tension with PCE was considerably lower at 0.160 dynes cm–1 but the solubility was reported to be 76 360 mg L–1. Both surfactant formulations were modified with salt to obtain densities >1.0 kg L–1, which was desired to avoid density override effects and to achieve complete flushing of the swept volume (Ramsburg and Pennell, 2001). The results showed that the Tween 80 application resulted in PCE solubilization with a removal percentage of only 53%; however, although the Aerosol MA-80 was able to remove 78%, the application resulted in downward PCE mobilization. An analysis based on the trapping number showed that this result was not unexpected. As a result of this study, Aerosol MA-80, although a great solubilizer, was removed as potential surfactant to clean up the targeted field site.
Conrad et al. (2002) conducted controlled experiments using Aerosol MA and Tween 80 surfactants to increase the solubility of TCE in sand packs designed to be representative of a fluvial depositional environment. The TCE was injected from a point source in a fully water-saturated system. The final distribution of the TCE revealed a series of descending pools. The Aerosol MA surfactant caused extreme reductions (almost 50-fold) in the DNAPL–water interfacial tension, followed by mobilization into low-permeability regions, worsening the remediation problem. Application of the Tween 80 surfactant, with a smaller interfacial tension reduction, resulted in modest, manageable mobilization since the liquid TCE did not migrate into low-permeability regions. The favorable solubilization properties of TCE in a Tween 80 solution allowed a recovery of almost 90%. Conrad et al. (2002) noted that the propensity to mobilize during surfactant flushing is greatly enhanced when DNAPL is held in pools.
Schaerlaekens and Feyen (2004) completed three experiments in which relatively small amounts of entrapped TCE were removed with a 2% Tween 80 solution using different flow rates. Only small amounts of TCE were injected to ensure that pooling would not occur at the bottom of the cell and that all TCE would be in the entrapped form. Similar to Saba et al. (2001), one of the goals of this work was to determine an applicable empirical mass transfer rate model for intermediate-scale experiments. The rate-limited solubilization process was simulated with the so-called NAPL code (Guarnaccia et al., 1997). The derived Gilland–Sherwood expression showed, not surprisingly, a strong correlation between the rate coefficient and the NAPL saturation and the Reynolds number. No attempts were made to apply the derived expression to other 2-D experiments. Schaerlaekens and Feyen (2004) compared their 2-D expression with a relationship derived for earlier column experiments using the same porous materials and chemical. Consistent with Saba and Illangasekare (2000) and Saba et al. (2001), they found that the expression derived for the 2-D cell predicts lower mass transfer rates than the model that was obtained through column experiments. Schaerlaekens and Feyen (2004) argued that the lower mass transfer rate in their 2-D systems could be attributed to bypassing of the DNAPL zone in a 2-D system due to a reduced permeability.
Single-well push–pull tests were conducted by Field et al. (1999, 2000). The experiments discussed in Field et al. (1999) were conducted to evaluate the ability of this test to characterize the enhanced solubilization of TCE by the surfactant DOWFAX (hexadecyl diphenyl oxide disulfonate). The experiments were performed in wedge-shaped cells to simulate the alternating radially divergent–convergent flow field in the vicinity of an injection–extraction well. Experiments with DOWFAX considerably increased the maximum concentration of the extracted solutions compared with flushings with water only; however, Field et al. (1999) also reported considerabe sinking of the surfactant solution due to solubilization of TCE and subsequent accumulation of DOWFAX and dissolved TCE at greater depth. No mobilization or sinking of liquid TCE was observed. Based on their results, Field et al. (1999) stated that, despite the dissolved plume sinking, push–pull tests can provide useful information on surfactant-enhanced solubilization of NAPLs in the subsurface.
Surfactant effectiveness for enhancing NAPL solubility has been evaluated in the laboratory for a wide range of system-specific conditions, such as the type and concentration of surfactants and cosolvents, and the valence and concentrations of electrolytes present in the applied solutions and in the solid phase (Edwards et al., 1991; Rouse et al., 1993; Valsaraj and Thibodeaux, 1989). Consequently, it is to be expected that cation exchange between the liquid and solid phase will play a role in the effectiveness of surfactant-enhanced NAPL recovery in the subsurface. This was the topic of further research by Field et al. (2000) in a push–pull (single well) flow cell experiment. They assessed the potential for cation exchange to adversely affect the phase behavior of the surfactant Aerosol MA 80-I (sodium dihexyl sulfosuccinate) and its solubilization of TCE in the subsurface. Based on batch experiments, they concluded that TCE solubility can either increase (at intermediate Ca2+ concentrations) or decrease (at high Ca2+ concentrations, either a Winsor Type II or III solution) in the surfactant solution they used. Their flow cell experiment showed that increasing concentrations of Ca2+ and Mg2+ resulted in the partitioning of sulfosuccinate into the entrapped TCE phase and smaller aqueous TCE concentrations than predicted from the solubilization isotherm. Field et al. (2000) also showed that quantities of exchangeable Ca2+ and Mg2+ could be substantially reduced by a 0.130 mol L–1 NaCl preflush, and restoring conservative sulfosuccinate transport, Winsor Type I phase behavior, and increased aqueous TCE concentrations. The simultaneous reduction of interfacial tensions, however, resulted in free TCE accumulation.
Alcohol Flushing
Alcohol flushing for environmental purposes, like surfactant flushing, is derived from an enhanced oil recovery technique mainly for crude oil recovery. Many alcohols are mutually miscible in both water and NAPL and possess a great ability to solubilize and mobilize NAPLs as pure phase by a combination of viscous and gravity forces, and via reduction of interfacial tensions (Grubb et al., 1997). When used in combination with surfactants (see combination schemes below), alcohols are sometimes referred to as cosolvents. Alcohols that are considered for remediation (e.g., ethanol, 1-propanol, and 2-propanol) are usually low-molecular-weight alcohols that are completely soluble in water. The fact that the specific gravities of LNAPLs and many alcohols are <1 provides a tendency for both to accumulate near the water table. Intermediate-scale experiments for LNAPL removal were reported by Grubb et al. (1997, 1998) and Palomino and Grubb (2004). When DNAPLs are being remediated with alcohols, concerns exist that the DNAPLs may mobilize and contaminate deeper regions previously not contaminated. Some of the earlier 1-D column experiments were conducted to promote DNAPL swelling, i.e., to have the alcohol partition into a DNAPL (e.g., Boyd, 1991; Brandes and Farley, 1993) to reduce its density. The experimental work of these researchers assessed the utility of alcohols for enhancing the mobilization of entrapped DNAPLs in 1-D columns through interfacial tension reduction and DNAPL swelling. Their investigations used isopropyl and tert-butyl alcohols in combination with the DNAPLs TCE and PCE. Downward migration during the upward flushing with alcohol–water mixtures was minimized because of the partitioning of alcohols into the DNAPLs. Most of the flow cell remediation experiments involving alcohol flushing of DNAPLs (Boyd et al., 2006; Grubb and Sitar, 1999; Lunn and Kueper, 1997, 1999a, 1999b; Roeder et al., 2001; Van Valkenburg and Annable, 2002) focused on mobilization reduction.
