Environ. Sci. Technol. 1997, 31, 3565-3572
Surfactant-Enhanced Remediation of a Trichloroethene-Contaminated Aquifer. 1. Transport of Triton X-100 J A M E S A . S M I T H , * ,† D I P A K S A H O O , † H E A T H E R M . M C L E L L A N , †,‡ A N D THOMAS E. IMBRIGIOTTA§ Program of Interdisciplinary Research in Contaminant Hydrogeology, Department of Civil Engineering, University of Virginia, Charlottesville, Virginia 22903-2442, and U.S. Geological Survey, 810 Bear Tavern Road, Suite 206, West Trenton, New Jersey 08628
Transport of a nonionic surfactant (Triton X-100) at aqueous concentrations less than 400 mg/L through a trichloroethene-contaminated sand-and-gravel aquifer at Picatinny Arsenal, NJ, has been studied through a series of laboratory and field experiments. In the laboratory, batch and column experiments were conducted to quantify the rate and amount of Triton X-100 sorption to the aquifer sediments. In the field, a 400 mg/L aqueous Triton X-100 solution was injected into the aquifer at a rate of 26.5 L/min for a 35-d period. The transport of Triton X-100 was monitored by sampling and analysis of groundwater at six locations surrounding the injection well. Equilibrium batch sorption experiments showed that Triton X-100 sorbs strongly and nonlinearly to the field soil with the sharpest inflection point of the isotherm occurring at an equilibrium aqueous Triton X-100 concentration close to critical micelle concentration. Batch, soil column, and field experimental data were analyzed with zero-, one-, and two-dimensional (respectively) transient solute transport models with either equilibrium or rate-limited sorption. These analyses reveal that Triton X-100 sorption to the aquifer solids is slow relative to advective and dispersive transport and that an equilibrium sorption model cannot simulate accurately the observed soil column and field data. Comparison of kinetic sorption parameters from batch, column, and field transport data indicate that both physical heterogeneities and Triton X-100 mass transfer between water and soil contribute to the kinetic transport effects.
Introduction Research over the past 10 yr has shown that the use of surfaceactive agents (surfactants) has the potential to increase the rate of remediation of ground water contaminated with relatively nonpolar organic pollutants (1-12). At aqueous concentrations above critical micelle concentration (cmc), surfactants can increase the apparent water solubility of organic pollutants (13, 14). When residual concentrations of non-aqueous-phase liquids (NAPLs) are present in a porous * Author to whom correspondence should be addressed; e-mail address:
[email protected]; telephone: (804)924-7991; fax: (804)982-2951. † University of Virginia. ‡ Present address: Woodward-Clyde, Inc., Blue Bell, PA. § U.S. Geological Survey.
S0013-936X(97)00314-3 CCC: $14.00
1997 American Chemical Society
medium, this solubility-enhancing effect can increase the rate of NAPL dissolution (1, 8, 9). For pollutants sorbed to soil, the solubility-enhancing effect of surfactants can increase the rate of pollutant desorption from soil to water (1-3, 5, 6, 15-19). Certain surfactants can also increase the masstransfer coefficient for pollutant desorption from soil to water, presumably by swelling the soil organic matter and lowering diffusional resistances of the solute in the soil organic matter (3, 5, 15). Over a limited range of concentrations near cmc, certain surfactants may sorb to natural soil and increase pollutant sorption (5, 6, 19). By reducing interfacial tensions, surfactants can mobilize NAPLs (1). Surfactants can also affect natural microbial populations in the subsurface (4, 12, 2022). Given the breadth and complexity of environmental applications of surfactants, a detailed understanding of the transport of surfactants in the subsurface is essential for the proper implementation of surfactant remediation technologies. This paper is the first of two publications addressing the results of a field test of surfactant-enhanced aquifer remediation at Picatinny Arsenal, NJ. Unlike previous field experiments, this experiment is designed to examine the effect of a relatively low concentration (less than 400 mg/L) of a nonionic surfactant (Triton X-100) on the rate of desorption of TCE from a contaminated water table aquifer currently undergoing pump-and-treat remediation. Despite extensive characterization of the field site over the past 10 yr, nonaqueous-phase TCE has not been detected in the aquifer. This work follows several other laboratory-based investigations that have shown that Triton X-100 can increase the soil water mass-transfer coefficient of organic pollutants at concentrations as low as 30 mg/L (3, 5, 15). This paper focuses on the transport of a nonionic surfactant, Triton X-100 through the unconfined, TCE-contaminated aquifer at the site. The primary objective of this work is to gain better understanding of surfactant transport under field conditions with a specific emphasis on quantification of the magnitude and rate of surfactant sorption to the aquifer solids. This information is a prerequisite to further studies of surfactant-enhanced aquifer remediation using relatively low concentrations of nonionic surfactants. The second paper in this series will address the effects of Triton X-100 on the desorption and transport of TCE in the aquifer. Description of Field Site. Picatinny Arsenal (Figure 1) is located in a narrow, glaciated valley in north-central New Jersey atop 50-65 m of stratified and unstratified drift, which in turn overlies a weathered bedrock surface. Three major hydrogeologic units encompassing the unconsolidated sediments at the site are (i) a 15-21 m thick unconfined sandand-gravel aquifer; (ii) an 8-21 m thick confining layer composed of fine sand, silt, and clay; and (iii) an 8-35 m thick confined sand-and-gravel aquifer (23). The water table is 2-4 m below land surface over most of the study area (24). The approximate direction of groundwater flow in the unconfined aquifer is from Building 24 to Green Pond Brook (Figure 1). The horizontal hydraulic conductivity, K, in the unconfined aquifer ranges from 0.01 to 0.13 cm/s, and average linear groundwater flow velocities range from 0.3 to 0.9 m/d (25, 26). From 1960-1989, Building 24 (Figure 1) was used for metal plating, cleaning, and degreasing operations. Trichloroethene was the primary solvent used for degreasing. Wastewater from these operations was discharged into two sandbottomed settling lagoons upgradient of the building (25). From 1973 to 1985, solvent vapors from a degreasing unit in Building 24 condensed into an improperly installed overflow pipe and were discharged into a dry well immediately
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downgradient of Building 24 (25). Because of these improper disposal practices, a plume of TCE-contaminated groundwater developed in the unconfined aquifer from Building 24 to Green Pond Brook (Figure 1). TCE concentrations as high as 44 mg/L have been measured in groundwater samples. In 1986, this site was selected by the U.S. Geological Survey as a national research site for the study of the fate, transport, and remediation of chlorinated solvents in groundwater. Because of this action, the site hydrogeology and contamination has been thoroughly characterized over the past decade (23-25, 27-30). In 1992, a pump-and-treat system consisting of five withdrawal wells and an air-stripping tower and activatedcarbon tanks was installed to remediate the contaminated groundwater. Based on the period from September 1992 (when the system was first activated) to February 1995, the pump-and-treat system had removed TCE from the aquifer at a rate of approximately 180 kg/yr (26). On the basis of the 1995 data alone, the rate of TCE withdrawal is only 70 kg/yr. This amount is small relative to the total mass of TCE purchased for use in Building 24 from 1974 to 1984 (approximately 67 000 kg) and comparable to the net transport of TCE vapor through the unsaturated zone to the atmosphere (50 kg/yr) (26) and the groundwater discharge of TCE into Green Pond Brook (50 kg/yr) (25). In addition, since pumpand-treat remediation began in 1992, there has been no significant decrease in groundwater TCE levels in the aquifer in areas more than 40 m from the withdrawal wells. The relative inefficiency of the pump-and-treat system has been attributed at least in part to the rate-limited desorption of TCE from the long-term contaminated aquifer solids (5, 25).
Materials and Methods Laboratory Experiments. Triton X-100, a nonionic heterogeneous octylphenol ethoxylate surfactant, was obtained from Union Carbide Chemical and Plastics Technology Corp. Potassium bromide (>99% purity), cobalt(II)nitrate hexahydrate (>98% purity), ammonium thiocyanate (>97.5% purity), sodium azide (>99% purity), and sodium chloride (>99% purity) were obtained from Aldrich Chemical Company. Spectrophotometric grade benzene (>99% purity) was obtained from Fisher Scientific. Soil from the field site was collected from depths of 9.5-16 m by hollow-stem augering with a split-spoon sampler during the installation of the seven monitoring and injection wells shown in Figure 1. These soil samples were composited and air-dried for 24 h for use in all laboratory batch and column experiments. The composite soil sample is a fine-to-medium sand with an organic carbon content of 0.08%. Batch sorption experiments were conducted in the laboratory to quantify the rate of Triton X-100 sorption to the field soil and the equilibrium distribution of Triton X-100 between the field soil and water. For the sorption rate experiments, 6 g of soil and 12 mL of an aqueous Triton X-100 solution at a concentration of 1086 mg/L were combined in 15-mL glass centrifuge tubes with Teflon-lined caps. The batch reactors were gently shaken in the dark at 20 °C until ready for sampling. After different time periods, duplicate reactors were removed from the shaker and centrifuged at 2000g for 30 min. The supernatant was analyzed to determine the concentration of Triton X-100 using a calibrated Fisher Scientific 20 surface tensiometer. The Triton X-100 quantification limit for the surface tensiometer is 10 mg/L. In addition to the above-described batch reactors, a series of “blank” reactors consisting of water and Triton X-100 but no soil were also prepared. These reactors were treated identically to the reactors with soil to determine if other processes such as sorption to glassware or biodegradation of surfactant contributed to the decrease in aqueous surfactant concentration versus time. Less than a 5% decrease in surfactant concentration in these batch reactors over time was observed.
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FIGURE 1. Map of the field site showing the location of study wells, pump-and-treat withdrawal wells, and region of trichloroethenecontaminated groundwater. Analysis of the kinetic batch experiments described above indicated that at least 90% of equilibrium was reached within approximately 96 h. Therefore, batch experiments to quantify the equilibrium Triton X-100 isotherm were allowed to equilibrate for a 96-h period. For these experiments, batch reactors were prepared in a similar fashion to the reactors described above, except that a range of aqueous solutions with different Triton X-100 concentrations was added to the individual batch reactors. After 96 h, all the reactors were centrifuged at 2000g for 30 min, and the supernatant was analyzed for Triton X-100 by surface tensiometry. The sorbed concentration of Triton X-100 was determined by difference. Two “blank” batch reactors containing water and Triton X-100 but no soil were treated identically to these tubes for quality control. Approximately 95% of the added surfactant was accounted for in these tubes at the conclusion of the 96-h equilibration period. For equilibrium aqueous surfactant concentrations greater than cmc (approximately 130 mg/L (31)), the supernatant was diluted to a concentration less than cmc prior to analysis by surface tensiometry. Triplicate column experiments were conducted using 30 cm long, 2.5 cm diameter glass columns packed with soil from the field site. Water without Triton X-100 was pumped through the columns (from bottom to top) with a Manostat peristaltic cassette pump at a rate of 15 ( 1 mL/h until steadystate flow conditions were reached. Next, an aqueous solution containing 200 mg/L sodium azide (to prevent microbial activity) and 117 mg/L Triton X-100 was pumped through the column at the same flow rate. Previous research has shown that sodium azide at this concentration does not affect TCE sorption (3, 5). Effluent samples (4 mL) from the column were collected over a 5.5-wk period (the time required to observe complete breakthrough of the surfactant), and the Triton X-100 concentration in the effluent samples was quantified over time by surface tensiometry.
TABLE 1. Hydraulic Conductivities As Determined from Slug Tests, Mean and Range of Measured Hydraulic Heads (Relative to Mean Sea Level) for May-December 1995, and Model-Predicted Hydraulic Heads for Seven Wells well no. 92-13 92-14 92-15 92-16 92-17 92-18 92-19
hydraulic mean measured hydraulic model-predicted conductivity hydraulic head head range hydraulic head (m/s) (m) (m) (m) 1.1 × 10-4 1.0 × 10-5 0.36 × 10-4 0.53 × 10-4 1.1 × 10-4 1.7 × 10-4 2.6 × 10-4
209.938 211.491 211.688 211.319 209.969 209.842 209.791
0.057 0.096 0.041 0.044 0.020 0.018 0.013
209.875 209.947 210.193 210.062 209.931 209.872 209.826
After surfactant breakthrough, a Br- tracer test was conducted. A 425 mg/L KBr solution was pumped through the column for 31 h. Effluent samples were collected and analyzed for Br- using an Orion specific ion electrode and a double junction reference electrode. The tracer test ended after several effluent samples were observed to have Brconcentrations approximately equal to the influent Brconcentration. At the completion of the test, the porous medium bulk density and porosity were determined gravimetrically, and their mean values were found to equal 1590 g/L and 0.42, respectively. Field Experiment. The transport of Triton X-100 at the field site was studied using a series of seven 5 cm diameter poly(vinyl chloride) injection and monitoring wells installed approximately in the center of the contaminant plume along two intersecting perpendicular lines as indicated in Figure 1. In this region, the unsaturated zone is approximately 3.3 m thick, and the saturated zone extends from a depth of 3.3 m to a depth of 16.2 m. Wells 92-14, 92-15, and 92-16 were used as injection wells and are screened from a depth of 9.7-16.2 m. Wells 92-13, 92-17, 92-18, and 92-19 are monitoring wells and are screened from a depth of 12.2-13.8 m. All the wells are separated by a horizontal distance of 3.25 m from its nearest-neighbor well with the exception of 92-13, which is located a distance of 9.75 m from well 92-15. The hydraulic conductivity of the study area was estimated by a series of slug tests. For each of the seven wells, the natural water level was instantaneously increased by the addition of a slug of water. The elevation of the water level in the well during recovery was recorded over time by pressure measurements in the well collected with a Geokon vibratingwire pressure transducer and a Campbell CR10 data logger. The resulting water level versus time data were analyzed according to Hvorslev’s method to determine the hydraulic conductivity of the porous medium surrounding the well casing (32). From May 23 to December 26, 1995, water was injected into wells 92-14, 92-15, and 92-16 at rates of 2.0, 26.5, and 15.0 L/min, respectively. These injection rates resulted in approximately equal steady-state hydraulic heads for wells 92-14 (211.28 m) and 92-16 (211.22 m) and a slightly higher hydraulic head for well 92-15 (211.54 m). The relatively low injection rate in well 92-14 needed to produce a hydraulic head similar to the heads in the other two injection wells is likely caused by some clogging of the well during its installation. This observation is supported by the slug test results, which showed that the hydraulic conductivity determined from well 92-14 was approximately 10 times less than the values determined for the other wells (Table 1). On June 3, 1995, 20 L of water containing 7.5 kg of KBr (250 g/L Br-) was also discharged into well 92-15 over a 20-min period. From June 28 to August 2, 1995, water discharged into injection well 92-15 contained Triton X-100 at a concentration of 400 mg/L. Hydraulic heads in all the wells were measured 3-4 times each week during the course of the field experiment.
No systematic changes in water table position were observed during this period. Groundwater samples were collected from the seven monitoring and injection wells and analyzed for one or more of the following constituents: Br-, TCE, Triton X-100. Groundwater samples were collected with a Grundfos RediFlo2 submersible pump and Teflon tubing using a low-flow purging technique similar to that described by Puls and Paul (33). For each well, the pump was lowered to a depth of 13 m, and groundwater was pumped to land surface at a rate of 1 L/min until one casing volume of water was removed, and the pH, specific conductance, and temperature of the groundwater stabilized. Water quality parameters were quantified with a YSI 3560 water quality monitoring system equipped with an in-line flow-through chamber and pH, specific conductance, and temperature probes. Following well purging, triplicate groundwater samples were collected in 40-mL amber glass sampling vials with Teflon-lined septum caps. Care was taken to ensure that no air bubbles were present in the vials after sample collection. The samples were stored at 5 °C in the dark until ready for analysis. All samples were analyzed for Triton X-100 or Br- within 3 d of sample collection. After sampling a well, the pump and tubing were washed with a 10:1 water:methanol mixture followed by a second washing with water. Samples of the final wash water were periodically collected for analysis to ensure that there was no cross contamination between wells. If the concentration of the solute in the wash blank was greater than 10% of the reported concentration in the sampled well, the data were discarded. In almost all cases, the solute concentration in the wash blank was less than 2% of the concentration in the groundwater sample. The concentration of Br- in groundwater samples was determined as described previously for column tracer experiments. The concentration of Triton X-100 in groundwater samples was determined by a spectrophotometric method (34, 35). A 10-mL sample of groundwater was mixed with 3-4 g of NaCl and 3 mL of ammonium cobalt thiocyanate reagent in a 50-mL centrifuge tube. The reagent was prepared by dissolving 62 g of ammonium thiocyanate and 28 g of cobalt nitrate hexahydrate in distilled water and diluting the mixture to 100 mL. The solution was shaken until the salt dissolved and was allowed to stand for 15 min. The surfactant mixture was extracted into 15 mL of benzene. The absorbance of the organic solution is measured with a Shimadzu UV1201S spectrophotometer at a wavelength of 320 nm. The spectrophotometer was zeroed with a blank water sample (e.g., a sample without Triton X-100) carried through the above-described procedure. Absorbance is linearly related to surfactant concentration, thereby facilitating calibration of the instrument with Triton X-100 standards. This method has a lower quantification limit (1 mg/L) than the surface tensiometric technique (10 mg/L) used for surfactant analyses in column experiments and was therefore chosen for analysis of the field data. For all water samples containing Triton X-100 in batch, column, and field experiments, there was no visible evidence of the formation of emulsions or “gelling” of the surfactant due to poor phase behavior, indicating that Triton X-100 existed only as dissolved monomers or micelles in solution. Solute Transport Simulations. To simulate the batch laboratory experimental data, a two-site sorption model was employed (36). The model assumes that sorption sites on the soil can be divided into two typesssites that are locally in equilibrium with the solute concentration in the aqueous phase (equilibrium sites) and sites that offer mass-transfer resistance to sorbing solutes (kinetic sites). The governing equations for the two types of sorption sites are as follows:
∂Se ∂C ab )F 2 ∂t ∂t (1 + bC)
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∂Sk abC ) k (1 - F) - Sk ∂t 1 + bC
[
]
(2)
The above rate equations assume that, at equilibrium, sorption can be described by a nonlinear Langmuir equation, e.g.
S ) Se + Sk )
abC 1 + bC
(3)
where S is the sorbed concentration of the solute (M/M); Se is the sorbed solute concentration for equilibrium sites (M/ M); Sk is the sorbed solute concentration for kinetic sites (M/ M); C is the solute concentration in the aqueous phase (M/ L3); F is the fraction of equilibrium sites; a and b are the Langmuir parameters (M/M) and (L3/M), respectively; k is the soil water mass-transfer coefficient (1/T); and t is the time (T). Equations 1-3 were solved numerically for C as a function of time. The Langmuir parameters a and b were determined by equilibrium batch sorption experiments, and the kinetic sorption parameters F and k were determined by calibration of the model to the batch kinetic sorption data. For simulation of laboratory column experiments and the field experiment under steady-state flow conditions, the governing equation used is given below:
[
1+F
]
Fab ∂C F ∂Sk + ) ∇‚(D∇C) - ∇‚(vC) + G 2 ∂t θ ∂t (1 + bC) (4)
where D is the dispersion tensor, v is the average groundwater velocity vector, F is the bulk density, θ is the porosity, G is an external supply term, and ∂Sk/∂t is defined in eq 2. For column experiments, the one-dimensional form of eq 4 was solved numerically for constant D and v and for G ) 0. The parameters D and v were determined from model calibration using Br- tracer data for each column experiment. The Langmuir isotherm parameters for Triton X-100 sorption were determined by batch sorption experiments described previously. The kinetic sorption parameters F and k in eq 2 were determined from model calibration using the observed surfactant concentration data. The initial and boundary conditions used for simulation of the column experimental data were as follows:
C(x,0) ) Se(x,0) ) Sk(x,0) ) 0
(C - Dv ∂C∂x )|
) Cin
(5) (6)
x)0
∂C(∞,t) ) finite ∂x
(7)
where Cin is the inflow solute concentration. For simulation of Triton X-100 transport at the field site, SUTRA (SaturatedUnsaturated TRAnsport), a finite-element flow and solutetransport model developed by the U.S. Geological Survey, was used (37). The code was modified to account for twosite kinetic sorption as formulated in eq 4. The modified code was tested against MSORB, an independently developed solute transport code that also accounts for two-site kinetic sorption (38) for transient one-dimensional solute transport problems, and comparison simulations yielded nearly identical results. Two-dimensional (vertically averaged) simulations were performed for steady-state groundwater flow and transient solute transport for a 21 by 36 m rectangular region (with boundaries parallel to either the monitoring or injection well transect) containing the injection and monitoring wells. Distances between node points ranged from 0.027 to 1.75 m, with the finest discretization established near injection well 92-15 where the sharpest solute concentration fronts occurred. For the flow model, constant hydraulic head boundary
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conditions were assumed for the upper (northwesterly) and lower (southeasterly) boundaries, and no-flow boundary conditions were assumed for the left and right (lateral) boundaries. The constant-head boundary values were determined by a linear extrapolation of the natural gradient determined from measurement of the hydraulic heads in wells 92-13, 92-15, 92-17, 92-18, and 92-19 (e.g., a linear regression line was fit to a plot of hydraulic head in the aforementioned wells versus location, and the linear fit was extrapolated to the boundaries of the simulation domain to determine the boundary heads). These extrapolated boundary heads agreed well with heads measured in wells located near the upper and lower simulation boundaries. For solute transport simulations, the initial surfactant concentration throughout the region was set equal to 0 mg/L. Zero concentration gradient boundary conditions were used for all four boundaries. Calibration of the flow model to the steady-state hydraulic heads in the monitoring wells resulted in a constant aquifer hydraulic conductivity of 1.3 × 10-2 cm/s. The porosity for all simulations is 0.28. The measured steady-state heads and the calibrated model’s predictions are given in Table 1. This hydraulic conductivity value did not provide good matches of the steady-state hydraulic heads in the injection wells. However, the calibrated hydraulic conductivity agreed well with the values determined from slug tests as part of this study (Table 1), with the hydraulic conductivity calibrated by the U.S. Geological Survey using a flow model for the entire region of contaminated groundwater and with the result of large-scale pump tests conducted previous to this study at Picatinny Arsenal (39). To accurately simulate the heads in the injection wells, zones of low permeability were required surrounding these wells. Since there was no physical evidence of these low-permeability zones in the field and since a single permeability value accurately simulated the observed heads in the monitoring well, the single value of 1.3 × 10-2 cm/s was used for all subsequent simulation. The data from the Br- tracer test were used to determine the longitudinal and transverse dispersivities. The Langmuir isotherm parameters for Triton X-100 sorption used in the simulations of surfactant transport were determined by the batch sorption experiments described previously. The kinetic sorption parameters F and k in eqs 2 and 4 were determined from model calibration using the observed surfactant concentration data from the monitoring wells. For all simulations, it was assumed that Triton X-100 was not being biologically or chemically transformed and that the oligomers in the heterogeneous surfactant did not fractionate during sorption to soil. All model fits were performed by trial and error, except for determination of dispersion coefficients and advection velocities in tracer experiments, which were determined by an optimization routine in the model CXTFIT (36).
Results The sorption isotherm in Figure 2 describes the equilibrium distribution of Triton X-100 between water and the field soil. The isotherm is distinctly nonlinear and is well-described by the Langmuir model (eq 3) with a ) 2.5 mg/g and b ) 0.015 L/mg. The sharpest curvature on the isotherm occurs at an aqueous concentration close to cmc (approximately 130 mg/L (31)), and this behavior is consistent with results reported elsewhere for Triton X-100 sorption (7, 16, 18, 19, 40). The isotherm data in Figure 2 could also be described by a twopart linear isotherm indicating strong, linear uptake of Triton X-100 at sub-cmc concentrations and essentially zero additional uptake of Triton X-100 at supra-cmc concentrations, although this approach would not be consistent with approaches used in previous investigations (7, 16, 18, 19, 40). Figure 3 describes the rate-limited uptake of Triton X-100 in batch sorption experiments. Equilibrium does not appear to be reached for approximately 96 h. The kinetic sorption
FIGURE 2. Equilibrium isotherm for Triton X-100 sorption from water to soil.
FIGURE 3. Aqueous Triton X-100 concentration versus time in soil water batch reactors.
TABLE 2. Values of Parameters F (Fraction of Equilibrium Sorption Sites) and k (Mass-Transfer Coefficient) for Simulations of Triton X-100 Sorption to Soil from Water for Laboratory Batch and Column Experiments and a Field Experimenta parameter
batch experiments
column experiments
field experiment
F k (s-1)
0.47 3.1 × 10-6
0.19 1.5 × 10-6
0.15 1.5 × 10-7
a
All experiments used the same soil type.
parameters F and k for the model fit of the data in Figure 3 are given in Table 2. Figure 4 presents the effluent concentrations of Triton X-100 from triplicate column experiments as a function of time. The break-through curves are approximately S-shaped, although there is considerable tailing, indicative of nonequilibrium (e.g., rate-limited) sorption. Two model fits are shown for each data set in Figure 4. The model fits indicated by the dashed lines are based on the assumption that a local equilibrium exists between the soil and water concentrations of Triton X-100 (e.g., F ) 1 in eqs 2 and 4). The model fits indicated by the solid lines do not assume a local sorption equilibrium (e.g., F < 1). The average kinetic sorption parameters (F and k) for the solid-line model fits of the data in Figure 4 are given in Table 2. The linear velocities (v) and dispersion coefficients (D) were determined by calibration of the model to the effluent concentrations of the nonsorbing Br- tracer (data not shown). Values of v and D for the column experiments are given in Table 3. The observed variability in the dispersion coefficient in Table 3 is likely caused by differences in the column packing. Figure 5 presents the results of the Br- tracer test at the field site. Measured and simulated tracer concentrations and model fits are given as a function of time for wells 92-17,
FIGURE 4. Triton X-100 effluent concentrations versus time and equilibrium and kinetic model fits for triplicate column experiments.
TABLE 3. Parameters Used To Simulate Transport of Triton X-100 through Triplicate Soil Columnsa column
v (cm/h)
D (cm2/h)
column
v (cm/h)
D (cm2/h)
A B
6.15 6.56
20.9 147
C mean
6.48 6.40
45.2 71.2
a The average velocities (v) and dispersion coefficients (D) were determined from Br- tracer tests and optimized model fits.
92-18, and 92-19. The longitudinal and tranverse dispersivities were calibrated to the field tracer data to obtain the model fit in Figure 5. Three values of longitudinal dispersivity (0.08 m, 0.24 m, and 1.04 m) were used to calibrate the model to the observed data with the magnitude of the longitudinal dispersivity increasing with distance from the injection well. The transverse dispersivity was assigned a value equal to onetenth the longitudinal dispersivity. Figure 6 presents the Triton X-100 concentration data in wells 92-13, 92-17, 92-18, and 92-19 as a function of time. Graph A in Figure 6 also gives the model fit to the data assuming that a local sorption equilibrium can be assumed. Only the model fit for well 92-17 is shown, as model fits for the other wells do not appear until times greater than 200 d. Graph B in Figure 6 gives the model fit for the surfactant data
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FIGURE 5. Concentration of Br- in groundwater samples collected from three wells as a function of time and model fit of the data.
FIGURE 6. Concentration of Triton X-100 in groundwater samples collected from four wells as a function of time and equilibrium (graph A) and kinetic (graph B) model fits of the data. assuming that sorption is rate limited and can be described by the two-site formulation.
Discussion The experimental data and analyses presented in the previous section indicate that Triton X-100 sorption to natural soil is rate-limited and cannot be accurately described by an equilibrium sorption model. This observation is true for laboratory batch and column experiments (Figures 3 and 4) and for the field experiments (Figure 6). The rate-limited sorption of Triton X-100 in laboratory experiments has been reported previously (7), but the comparison of sorption rates for a single surfactant and soil under different experimental conditions (e.g., batch, column, and field) provides additional insight into the causes of rate-limited sorption. Table 2 summarizes the kinetic rate parameters, F and k, for each type of experiment. Comparison of these values suggests that mass-transfer rate limitations are occurring at both the particle scale and at the porous-media scale. As the type of experiment changes from batch to column to field, the parameter F, which equals the fraction of equilibrium sites, decreases; the parameter k, which is the mass-transfer coefficient between soil and water, also decreases. Therefore,
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the sorption rate limitations become greater as the experiment type changes from batch to column to field. In the batch system, the soil water slurry is being mechanically mixed during incubation. Therefore, any rate-limited sorption is caused at the particle scale by either diffusion into the soil organic matter or diffusion into intraparticle pore channels. In the column experiments, heterogeneities and dead-end pores are likely introduced despite attempts to pack the soil uniformly into the soil column. Therefore, decreases in F and k as the experimental procedure is changed from batch to column is indicative of a “dual porosity” system. Part of the pore space is filled with mobile water that can advect the Triton X-100 (e.g., the effective porosity) whereas another part of the pore space contains stagnant water. Transfer of the solute between the mobile and immobile regions can only occur by molecular diffusion. At the field scale, heterogeneities become more pronounced than those found in the column experiment. In the field, zones of significantly varying permeabilities may exist in the aquifer in the region of the surfactant transport experiment. Furthermore, chemical (sorption) heterogeneities may also contribute to the strong kinetic effects observed in the field experiment (41). Chemical heterogeneities arise from soil with varying sorptive properties (e.g., different organic carbon contents, intraparticle porosity, etc.). Although soil samples used in laboratory experiments were collected during the drilling of the monitoring and injection wells for the field test, it is likely that the chemical heterogeneities encountered by the surfactant as it is transported through the aquifer are greater than those encountered by the surfactant in the soil column experiments. These physical and chemical heterogeneities require further decreases in the kinetic sorption parameters F and k to allow adequate model fits of the column and field data. It should be noted that if physical heterogeneities are indeed contributing to the decreasing values of F and k in the column and field experiments relative to the batch experiments, the meaning of these two parameters becomes ambiguous. For example, considering only physical heterogeneities, a more appropriate mathematical definition of the parameters would be the fraction of mobile water (F) and the mass-transfer coefficient for solute transport between mobile and immobile water (k). Because both rate-limited sorption and physical heterogeneities are affecting the transport of Triton X-100, a more physically realistic model of the field system may have to consider three phases: soil, mobile water, and immobile water. Mass-transfer limitations would exist between all three of these phases. The effects of field-scale physical heterogeneities are also evident from the Br- tracer-test data. Although Br- is generally considered to be a conservative (nonsorbing) tracer, it was difficult to obtain a precise match of the field tracer data with the solute-transport model and a constant value of the dispersivity (Figure 5). Review of Figure 5 shows that the peak Br- concentrations in wells 92-18 and 92-19 occur almost simultaneously and all of the peaks are nonsymmetric with extended “tails”. Cross-sectional maps of the field site developed by the U.S. Geological Survey (27) show several layers of lower permeability silty sand interspersed throughout the sand-and-gravel aquifer, offering further evidence of physical heterogeneities in the study area. To obtain the model fits shown in Figure 5, the dispersivity had to be increased with longitudinal distance from the injection well. This is equivalent to increasing the dispersivity for the entire model domain over time (e.g., as the surfactant pulse moves downgradient, larger dispersivity values are needed to account for the larger range of heterogeneities it encounters) and is consistent with the recognized observation that dispersivities increase with the scale of the problem (42-45). Therefore, it is apparent that both particle scale and porous-media scale diffusion processes are contributing to the rate-limited mass transfer of Triton X-100 in the aquifer at Picatinny Arsenal.
There may be two additional aspects of the experimental design that influence the tracer and surfactant transport experiments. First, the relatively high concentration of the tracer solution injected into the aquifer may have caused density-dependent flow near the injection well. This would result in greater “sinking” of the Br- plume than the surfactant plume because of its slightly greater density. Second, simulation of solute transport is performed with a twodimensional, vertically averaged model. Given that the monitoring and injection wells are not screened over the entire aquifer, it is possible that there were significant vertical concentration gradients in the aquifer during the field experiment. Both these aspects are likely not problematic because of the expected large extent of vertical mixing in the vicinity of the injection well casing caused by the large, induced vertical hydraulic gradient relative to the horizontal gradient. Therefore, it is likely that this vertical mixing has a much more significant effect on the tracer transport than density effects and strengthens the assumption of vertical averaging for both the tracer and the surfactant. Furthermore, if signficant vertical concentration gradients exist, these are present for both the tracer and the surfactant, and the modelcalibrated dispersivity accounts for the information loss due to vertical averaging. The rate limitations observed in this research at the laboratory and field scale have many important implications for surfactant-based remediation technologies. First, because of the observed rate-limited sorption, Triton X-100 that is injected into the subsurface will be transported significantly farther in a given time period than a local-equilibrium sorption model would predict. For example, the peak surfactant concentration in well 92-17 predicted by the local-equilibrium sorption model in Figure 6 appears at a time of about 120 d, whereas the actual surfactant peak concentration is observed around 40 d. This is probably a beneficial effect at the start of a surfactant remediation operation because it allows the surfactant to be transported greater distances in a given time period. Conversely, the rate-limited sorption may be problematic at the conclusion of a surfactant remediation operation if the surfactant must be removed from the subsurface. Because of the sorption rate limitations, it will require longer time periods to remove the surfactant from the aquifer than if sorption was governed by a local equilibrium. Compared to anionic surfactants and some other nonionic surfactants, Triton X-100 sorbs relatively strongly to the field soil (Figure 2), particularly for equilibrium aqueous concentrations less than cmc. This strong sorption would increase material costs for engineered surfactant remediation schemes, since a large fraction of the surfactant will be associated with the soil phase. Because of the flattening of the sorption isotherm at concentrations above cmc, this may be less problematic at concentrations well above cmc. However, the strong sorption of Triton X-100 at sub-cmc to soil may be the reason for its ability to increase the rate of desorption of other sorbed organic pollutants such as trichloroethene and carbon tetrachloride (3, 5). Interaction of the surfactant with the soil organic matter reduces the diffusional resistances of the soil organic matter and facilitates the outward diffusion of sorbed organic solutes into the bulk solution (3, 5).
Acknowledgments This research has been supported by the National Center for Environmental Research and Quality Assurance (NCERQA) of the U.S. Environmental Protection Agency and the Toxic Substances Hydrology Program of the U.S. Geological Survey. Part of the computational support for this research was obtained from a grant from the IBM Environmental Research Program. However, any opinions, findings, and conclusions or recommendations expressed in this material are those of the authors and do not necessarily reflect the views of the IBM Corporation.
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Received for review April 7, 1997. Revised manuscript received August 15, 1997. Accepted August 18, 1997.X ES970314V X
Abstract published in Advance ACS Abstracts, October 1, 1997.
