surfactant-enhanced remediation of a trichloroethene-contaminated aquifer. 1. transport of triton...

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Surfactant-Enhanced Remediation of a Trichloroethene-Contaminated Aquifer. 1. Transport of Triton X-100 JAMES A. SMITH,* DIPAK SAHOO, ² HEATHER M. MCLELLAN, ²,‡ AND 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 surface- active 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 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 mass- transfer 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, 20- 22). 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 technolo- gies. 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, non- aqueous-phase TCE has not been detected in the aquifer. This work follows several other laboratory-based investiga- tions 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 sedi- ments at the site are (i) a 15-21 m thick unconfined sand- and-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 sand- bottomed 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 * Author to whom correspondence should be addressed; e-mail address: [email protected]; telephone: (804)924-7991; fax: (804)- 982-2951. ² University of Virginia. Present address: Woodward-Clyde, Inc., Blue Bell, PA. § U.S. Geological Survey. Environ. Sci. Technol. 1997, 31, 3565-3572 S0013-936X(97)00314-3 CCC: $14.00 1997 American Chemical Society VOL. 31, NO. 12, 1997 / ENVIRONMENTAL SCIENCE & TECHNOLOGY 9 3565

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Surfactant-Enhanced Remediation ofa Trichloroethene-ContaminatedAquifer. 1. Transport of TritonX-100J 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 DT H O M A S E . I M B R I G I O T T A §

Program of Interdisciplinary Research in ContaminantHydrogeology, Department of Civil Engineering, University ofVirginia, Charlottesville, Virginia 22903-2442, andU.S. Geological Survey, 810 Bear Tavern Road, Suite 206,West Trenton, New Jersey 08628

Transport of a nonionic surfactant (Triton X-100) at aqueousconcentrations less than 400 mg/L through atrichloroethene-contaminated sand-and-gravel aquifer atPicatinny Arsenal, NJ, has been studied through a series oflaboratory and field experiments. In the laboratory, batchand column experiments were conducted to quantify therate and amount of Triton X-100 sorption to the aquifersediments. In the field, a 400 mg/L aqueous Triton X-100solution was injected into the aquifer at a rate of 26.5L/min for a 35-d period. The transport of Triton X-100 wasmonitored by sampling and analysis of groundwater atsix locations surrounding the injection well. Equilibriumbatch sorption experiments showed that Triton X-100 sorbsstrongly and nonlinearly to the field soil with the sharpestinflection point of the isotherm occurring at an equilibriumaqueous Triton X-100 concentration close to critical micelleconcentration. Batch, soil column, and field experimentaldata were analyzed with zero-, one-, and two-dimensional(respectively) transient solute transport models with eitherequilibrium or rate-limited sorption. These analyses revealthat Triton X-100 sorption to the aquifer solids is slow relativeto advective and dispersive transport and that an equilibriumsorption model cannot simulate accurately the observedsoil column and field data. Comparison of kinetic sorptionparameters from batch, column, and field transport dataindicate that both physical heterogeneities and Triton X-100mass transfer between water and soil contribute to thekinetic transport effects.

IntroductionResearch over the past 10 yr has shown that the use of surface-active agents (surfactants) has the potential to increase therate of remediation of ground water contaminated withrelatively nonpolar organic pollutants (1-12). At aqueousconcentrations above critical micelle concentration (cmc),surfactants can increase the apparent water solubility oforganic pollutants (13, 14). When residual concentrations ofnon-aqueous-phase liquids (NAPLs) are present in a porous

medium, this solubility-enhancing effect can increase therate of NAPL dissolution (1, 8, 9). For pollutants sorbed tosoil, the solubility-enhancing effect of surfactants can increasethe rate of pollutant desorption from soil to water (1-3, 5,6, 15-19). Certain surfactants can also increase the mass-transfer coefficient for pollutant desorption from soil to water,presumably by swelling the soil organic matter and loweringdiffusional 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 increasepollutant sorption (5, 6, 19). By reducing interfacial tensions,surfactants can mobilize NAPLs (1). Surfactants can also affectnatural microbial populations in the subsurface (4, 12, 20-22).

Given the breadth and complexity of environmentalapplications of surfactants, a detailed understanding of thetransport of surfactants in the subsurface is essential for theproper implementation of surfactant remediation technolo-gies. This paper is the first of two publications addressingthe results of a field test of surfactant-enhanced aquiferremediation at Picatinny Arsenal, NJ. Unlike previous fieldexperiments, this experiment is designed to examine the effectof a relatively low concentration (less than 400 mg/L) of anonionic surfactant (Triton X-100) on the rate of desorptionof TCE from a contaminated water table aquifer currentlyundergoing pump-and-treat remediation. Despite extensivecharacterization of the field site over the past 10 yr, non-aqueous-phase TCE has not been detected in the aquifer.This work follows several other laboratory-based investiga-tions that have shown that Triton X-100 can increase the soilwater mass-transfer coefficient of organic pollutants atconcentrations as low as 30 mg/L (3, 5, 15). This paper focuseson the transport of a nonionic surfactant, Triton X-100 throughthe unconfined, TCE-contaminated aquifer at the site. Theprimary objective of this work is to gain better understandingof surfactant transport under field conditions with a specificemphasis on quantification of the magnitude and rate ofsurfactant sorption to the aquifer solids. This information isa prerequisite to further studies of surfactant-enhancedaquifer remediation using relatively low concentrations ofnonionic surfactants. The second paper in this series willaddress the effects of Triton X-100 on the desorption andtransport of TCE in the aquifer.

Description of Field Site. Picatinny Arsenal (Figure 1) islocated in a narrow, glaciated valley in north-central NewJersey atop 50-65 m of stratified and unstratified drift, whichin turn overlies a weathered bedrock surface. Three majorhydrogeologic units encompassing the unconsolidated sedi-ments at the site are (i) a 15-21 m thick unconfined sand-and-gravel aquifer; (ii) an 8-21 m thick confining layercomposed of fine sand, silt, and clay; and (iii) an 8-35 mthick confined sand-and-gravel aquifer (23). The water tableis 2-4 m below land surface over most of the study area (24).The approximate direction of groundwater flow in theunconfined aquifer is from Building 24 to Green Pond Brook(Figure 1). The horizontal hydraulic conductivity, K, in theunconfined aquifer ranges from 0.01 to 0.13 cm/s, and averagelinear groundwater flow velocities range from 0.3 to 0.9 m/d(25, 26).

From 1960-1989, Building 24 (Figure 1) was used for metalplating, cleaning, and degreasing operations. Trichloroethenewas the primary solvent used for degreasing. Wastewaterfrom these operations was discharged into two sand-bottomed settling lagoons upgradient of the building (25).From 1973 to 1985, solvent vapors from a degreasing unit inBuilding 24 condensed into an improperly installed overflowpipe and were discharged into a dry well immediately

* Author to whom correspondence should be addressed; e-mailaddress: [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.

Environ. Sci. Technol. 1997, 31, 3565-3572

S0013-936X(97)00314-3 CCC: $14.00 1997 American Chemical Society VOL. 31, NO. 12, 1997 / ENVIRONMENTAL SCIENCE & TECHNOLOGY 9 3565

downgradient of Building 24 (25). Because of these improperdisposal practices, a plume of TCE-contaminated ground-water developed in the unconfined aquifer from Building 24to Green Pond Brook (Figure 1). TCE concentrations as highas 44 mg/L have been measured in groundwater samples. In1986, this site was selected by the U.S. Geological Survey asa 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 contamina-tion has been thoroughly characterized over the past decade(23-25, 27-30).

In 1992, a pump-and-treat system consisting of fivewithdrawal wells and an air-stripping tower and activated-carbon tanks was installed to remediate the contaminatedgroundwater. Based on the period from September 1992(when the system was first activated) to February 1995, thepump-and-treat system had removed TCE from the aquiferat a rate of approximately 180 kg/yr (26). On the basis of the1995 data alone, the rate of TCE withdrawal is only 70 kg/yr.This amount is small relative to the total mass of TCEpurchased for use in Building 24 from 1974 to 1984 (ap-proximately 67 000 kg) and comparable to the net transportof TCE vapor through the unsaturated zone to the atmosphere(50 kg/yr) (26) and the groundwater discharge of TCE intoGreen Pond Brook (50 kg/yr) (25). In addition, since pump-and-treat remediation began in 1992, there has been nosignificant decrease in groundwater TCE levels in the aquiferin areas more than 40 m from the withdrawal wells. Therelative inefficiency of the pump-and-treat system has beenattributed at least in part to the rate-limited desorption ofTCE from the long-term contaminated aquifer solids (5, 25).

Materials and MethodsLaboratory Experiments. Triton X-100, a nonionic hetero-geneous octylphenol ethoxylate surfactant, was obtained fromUnion Carbide Chemical and Plastics Technology Corp.Potassium bromide (>99% purity), cobalt(II)nitrate hexahy-drate (>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 ob-tained from Fisher Scientific. Soil from the field site wascollected from depths of 9.5-16 m by hollow-stem augeringwith a split-spoon sampler during the installation of the sevenmonitoring and injection wells shown in Figure 1. These soilsamples were composited and air-dried for 24 h for use in alllaboratory batch and column experiments. The compositesoil sample is a fine-to-medium sand with an organic carboncontent of 0.08%.

Batch sorption experiments were conducted in the labo-ratory to quantify the rate of Triton X-100 sorption to thefield soil and the equilibrium distribution of Triton X-100between the field soil and water. For the sorption rateexperiments, 6 g of soil and 12 mL of an aqueous Triton X-100solution at a concentration of 1086 mg/L were combined in15-mL glass centrifuge tubes with Teflon-lined caps. Thebatch reactors were gently shaken in the dark at 20 °C untilready for sampling. After different time periods, duplicatereactors were removed from the shaker and centrifuged at2000g for 30 min. The supernatant was analyzed to determinethe concentration of Triton X-100 using a calibrated FisherScientific 20 surface tensiometer. The Triton X-100 quan-tification limit for the surface tensiometer is 10 mg/L. Inaddition to the above-described batch reactors, a series of“blank” reactors consisting of water and Triton X-100 but nosoil were also prepared. These reactors were treated identi-cally to the reactors with soil to determine if other processessuch as sorption to glassware or biodegradation of surfactantcontributed to the decrease in aqueous surfactant concen-tration versus time. Less than a 5% decrease in surfactantconcentration in these batch reactors over time was observed.

Analysis of the kinetic batch experiments described aboveindicated that at least 90% of equilibrium was reached withinapproximately 96 h. Therefore, batch experiments to quantifythe equilibrium Triton X-100 isotherm were allowed toequilibrate for a 96-h period. For these experiments, batchreactors were prepared in a similar fashion to the reactorsdescribed above, except that a range of aqueous solutionswith different Triton X-100 concentrations was added to theindividual batch reactors. After 96 h, all the reactors werecentrifuged at 2000g for 30 min, and the supernatant wasanalyzed for Triton X-100 by surface tensiometry. The sorbedconcentration of Triton X-100 was determined by difference.Two “blank” batch reactors containing water and Triton X-100but no soil were treated identically to these tubes for qualitycontrol. Approximately 95% of the added surfactant wasaccounted for in these tubes at the conclusion of the 96-hequilibration period. For equilibrium aqueous surfactantconcentrations greater than cmc (approximately 130 mg/L(31)), the supernatant was diluted to a concentration lessthan cmc prior to analysis by surface tensiometry.

Triplicate column experiments were conducted using 30cm long, 2.5 cm diameter glass columns packed with soilfrom the field site. Water without Triton X-100 was pumpedthrough the columns (from bottom to top) with a Manostatperistaltic cassette pump at a rate of 15 ( 1 mL/h until steady-state flow conditions were reached. Next, an aqueous solutioncontaining 200 mg/L sodium azide (to prevent microbialactivity) and 117 mg/L Triton X-100 was pumped throughthe column at the same flow rate. Previous research hasshown that sodium azide at this concentration does not affectTCE sorption (3, 5). Effluent samples (4 mL) from the columnwere collected over a 5.5-wk period (the time required toobserve complete breakthrough of the surfactant), and theTriton X-100 concentration in the effluent samples wasquantified over time by surface tensiometry.

FIGURE 1. Map of the field site showing the location of study wells,pump-and-treat withdrawal wells, and region of trichloroethene-contaminated groundwater.

3566 9 ENVIRONMENTAL SCIENCE & TECHNOLOGY / VOL. 31, NO. 12, 1997

After surfactant breakthrough, a Br- tracer test wasconducted. A 425 mg/L KBr solution was pumped throughthe column for 31 h. Effluent samples were collected andanalyzed for Br- using an Orion specific ion electrode and adouble junction reference electrode. The tracer test endedafter several effluent samples were observed to have Br-

concentrations approximately equal to the influent Br-

concentration. At the completion of the test, the porousmedium bulk density and porosity were determined gravi-metrically, and their mean values were found to equal 1590g/L and 0.42, respectively.

Field Experiment. The transport of Triton X-100 at thefield site was studied using a series of seven 5 cm diameterpoly(vinyl chloride) injection and monitoring wells installedapproximately in the center of the contaminant plume alongtwo intersecting perpendicular lines as indicated in Figure 1.In this region, the unsaturated zone is approximately 3.3 mthick, and the saturated zone extends from a depth of 3.3 mto a depth of 16.2 m. Wells 92-14, 92-15, and 92-16 were usedas injection wells and are screened from a depth of 9.7-16.2m. Wells 92-13, 92-17, 92-18, and 92-19 are monitoring wellsand are screened from a depth of 12.2-13.8 m. All the wellsare separated by a horizontal distance of 3.25 m from itsnearest-neighbor well with the exception of 92-13, which islocated a distance of 9.75 m from well 92-15.

The hydraulic conductivity of the study area was estimatedby a series of slug tests. For each of the seven wells, thenatural water level was instantaneously increased by theaddition of a slug of water. The elevation of the water levelin the well during recovery was recorded over time by pressuremeasurements in the well collected with a Geokon vibrating-wire pressure transducer and a Campbell CR10 data logger.The resulting water level versus time data were analyzedaccording to Hvorslev’s method to determine the hydraulicconductivity of the porous medium surrounding the wellcasing (32).

From May 23 to December 26, 1995, water was injectedinto wells 92-14, 92-15, and 92-16 at rates of 2.0, 26.5, and15.0 L/min, respectively. These injection rates resulted inapproximately equal steady-state hydraulic heads for wells92-14 (211.28 m) and 92-16 (211.22 m) and a slightly higherhydraulic head for well 92-15 (211.54 m). The relatively lowinjection rate in well 92-14 needed to produce a hydraulichead similar to the heads in the other two injection wells islikely caused by some clogging of the well during itsinstallation. This observation is supported by the slug testresults, which showed that the hydraulic conductivity de-termined from well 92-14 was approximately 10 times lessthan the values determined for the other wells (Table 1). OnJune 3, 1995, 20 L of water containing 7.5 kg of KBr (250 g/LBr-) was also discharged into well 92-15 over a 20-min period.From June 28 to August 2, 1995, water discharged into injectionwell 92-15 contained Triton X-100 at a concentration of 400mg/L. Hydraulic heads in all the wells were measured 3-4times each week during the course of the field experiment.

No systematic changes in water table position were observedduring this period.

Groundwater samples were collected from the sevenmonitoring and injection wells and analyzed for one or moreof the following constituents: Br-, TCE, Triton X-100.Groundwater samples were collected with a Grundfos Redi-Flo2 submersible pump and Teflon tubing using a low-flowpurging technique similar to that described by Puls and Paul(33). For each well, the pump was lowered to a depth of 13m, and groundwater was pumped to land surface at a rateof 1 L/min until one casing volume of water was removed,and the pH, specific conductance, and temperature of thegroundwater stabilized. Water quality parameters werequantified with a YSI 3560 water quality monitoring systemequipped with an in-line flow-through chamber and pH,specific conductance, and temperature probes. Followingwell purging, triplicate groundwater samples were collectedin 40-mL amber glass sampling vials with Teflon-lined septumcaps. Care was taken to ensure that no air bubbles werepresent in the vials after sample collection. The samples werestored at 5 °C in the dark until ready for analysis. All sampleswere analyzed for Triton X-100 or Br- within 3 d of samplecollection. After sampling a well, the pump and tubing werewashed with a 10:1 water:methanol mixture followed by asecond washing with water. Samples of the final wash waterwere periodically collected for analysis to ensure that therewas no cross contamination between wells. If the concen-tration of the solute in the wash blank was greater than 10%of the reported concentration in the sampled well, the datawere discarded. In almost all cases, the solute concentrationin the wash blank was less than 2% of the concentration inthe groundwater sample.

The concentration of Br- in groundwater samples wasdetermined as described previously for column tracer experi-ments. The concentration of Triton X-100 in groundwatersamples was determined by a spectrophotometric method(34, 35). A 10-mL sample of groundwater was mixed with3-4 g of NaCl and 3 mL of ammonium cobalt thiocyanatereagent in a 50-mL centrifuge tube. The reagent was preparedby dissolving 62 g of ammonium thiocyanate and 28 g ofcobalt nitrate hexahydrate in distilled water and diluting themixture to 100 mL. The solution was shaken until the saltdissolved and was allowed to stand for 15 min. The surfactantmixture was extracted into 15 mL of benzene. The absorbanceof the organic solution is measured with a Shimadzu UV-1201S spectrophotometer at a wavelength of 320 nm. Thespectrophotometer was zeroed with a blank water sample(e.g., a sample without Triton X-100) carried through theabove-described procedure. Absorbance is linearly relatedto surfactant concentration, thereby facilitating calibrationof the instrument with Triton X-100 standards. This methodhas a lower quantification limit (1 mg/L) than the surfacetensiometric technique (10 mg/L) used for surfactant analysesin column experiments and was therefore chosen for analysisof the field data. For all water samples containing TritonX-100 in batch, column, and field experiments, there was novisible evidence of the formation of emulsions or “gelling” ofthe surfactant due to poor phase behavior, indicating thatTriton X-100 existed only as dissolved monomers or micellesin solution.

Solute Transport Simulations. To simulate the batchlaboratory experimental data, a two-site sorption model wasemployed (36). The model assumes that sorption sites onthe soil can be divided into two typesssites that are locallyin equilibrium with the solute concentration in the aqueousphase (equilibrium sites) and sites that offer mass-transferresistance to sorbing solutes (kinetic sites). The governingequations for the two types of sorption sites are as follows:

TABLE 1. Hydraulic Conductivities As Determined from SlugTests, Mean and Range of Measured Hydraulic Heads(Relative to Mean Sea Level) for May-December 1995, andModel-Predicted Hydraulic Heads for Seven Wells

wellno.

hydraulicconductivity

(m/s)

mean measuredhydraulic head

(m)

hydraulichead range

(m)

model-predictedhydraulic head

(m)

92-13 1.1 × 10-4 209.938 0.057 209.87592-14 1.0 × 10-5 211.491 0.096 209.94792-15 0.36 × 10-4 211.688 0.041 210.19392-16 0.53 × 10-4 211.319 0.044 210.06292-17 1.1 × 10-4 209.969 0.020 209.93192-18 1.7 × 10-4 209.842 0.018 209.87292-19 2.6 × 10-4 209.791 0.013 209.826

∂Se

∂t) F

ab

(1 + bC)2

∂C∂t

(1)

VOL. 31, NO. 12, 1997 / ENVIRONMENTAL SCIENCE & TECHNOLOGY 9 3567

The above rate equations assume that, at equilibrium, sorptioncan be described by a nonlinear Langmuir equation, e.g.

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 theLangmuir parameters (M/M) and (L3/M), respectively; k is thesoil water mass-transfer coefficient (1/T); and t is the time(T). Equations 1-3 were solved numerically for C as a functionof time. The Langmuir parameters a and b were determinedby equilibrium batch sorption experiments, and the kineticsorption parameters F and k were determined by calibrationof the model to the batch kinetic sorption data.

For simulation of laboratory column experiments and thefield experiment under steady-state flow conditions, thegoverning equation used is given below:

where D is the dispersion tensor, v is the average groundwatervelocity vector, F is the bulk density, θ is the porosity, G is anexternal supply term, and ∂Sk/∂t is defined in eq 2. For columnexperiments, the one-dimensional form of eq 4 was solvednumerically for constant D and v and for G ) 0. Theparameters D and v were determined from model calibrationusing Br- tracer data for each column experiment. TheLangmuir isotherm parameters for Triton X-100 sorption weredetermined by batch sorption experiments described previ-ously. The kinetic sorption parameters F and k in eq 2 weredetermined from model calibration using the observedsurfactant concentration data. The initial and boundaryconditions used for simulation of the column experimentaldata were as follows:

where Cin is the inflow solute concentration. For simulationof Triton X-100 transport at the field site, SUTRA (Saturated-Unsaturated TRAnsport), a finite-element flow and solute-transport model developed by the U.S. Geological Survey,was used (37). The code was modified to account for two-site kinetic sorption as formulated in eq 4. The modifiedcode was tested against MSORB, an independently developedsolute transport code that also accounts for two-site kineticsorption (38) for transient one-dimensional solute transportproblems, and comparison simulations yielded nearly identi-cal results. Two-dimensional (vertically averaged) simulationswere performed for steady-state groundwater flow andtransient solute transport for a 21 by 36 m rectangular region(with boundaries parallel to either the monitoring or injectionwell 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 well92-15 where the sharpest solute concentration fronts occurred.For the flow model, constant hydraulic head boundary

conditions were assumed for the upper (northwesterly) andlower (southeasterly) boundaries, and no-flow boundaryconditions were assumed for the left and right (lateral)boundaries. The constant-head boundary values were de-termined by a linear extrapolation of the natural gradientdetermined from measurement of the hydraulic heads in wells92-13, 92-15, 92-17, 92-18, and 92-19 (e.g., a linear regressionline was fit to a plot of hydraulic head in the aforementionedwells versus location, and the linear fit was extrapolated tothe boundaries of the simulation domain to determine theboundary heads). These extrapolated boundary heads agreedwell with heads measured in wells located near the upperand lower simulation boundaries. For solute transportsimulations, the initial surfactant concentration throughoutthe region was set equal to 0 mg/L. Zero concentrationgradient boundary conditions were used for all four bound-aries.

Calibration of the flow model to the steady-state hydraulicheads in the monitoring wells resulted in a constant aquiferhydraulic conductivity of 1.3 × 10-2 cm/s. The porosity forall simulations is 0.28. The measured steady-state heads andthe calibrated model’s predictions are given in Table 1. Thishydraulic conductivity value did not provide good matchesof the steady-state hydraulic heads in the injection wells.However, the calibrated hydraulic conductivity agreed wellwith the values determined from slug tests as part of thisstudy (Table 1), with the hydraulic conductivity calibrated bythe U.S. Geological Survey using a flow model for the entireregion of contaminated groundwater and with the result oflarge-scale pump tests conducted previous to this study atPicatinny Arsenal (39). To accurately simulate the heads inthe injection wells, zones of low permeability were requiredsurrounding these wells. Since there was no physical evidenceof these low-permeability zones in the field and since a singlepermeability value accurately simulated the observed headsin the monitoring well, the single value of 1.3 × 10-2 cm/swas used for all subsequent simulation. The data from theBr- tracer test were used to determine the longitudinal andtransverse dispersivities. The Langmuir isotherm parametersfor Triton X-100 sorption used in the simulations of surfactanttransport were determined by the batch sorption experimentsdescribed previously. The kinetic sorption parameters F andk in eqs 2 and 4 were determined from model calibrationusing the observed surfactant concentration data from themonitoring wells. For all simulations, it was assumed thatTriton X-100 was not being biologically or chemicallytransformed and that the oligomers in the heterogeneoussurfactant did not fractionate during sorption to soil. Allmodel fits were performed by trial and error, except fordetermination of dispersion coefficients and advectionvelocities in tracer experiments, which were determined byan optimization routine in the model CXTFIT (36).

ResultsThe sorption isotherm in Figure 2 describes the equilibriumdistribution of Triton X-100 between water and the field soil.The isotherm is distinctly nonlinear and is well-described bythe Langmuir model (eq 3) with a ) 2.5 mg/g and b ) 0.015L/mg. The sharpest curvature on the isotherm occurs at anaqueous concentration close to cmc (approximately 130 mg/L(31)), and this behavior is consistent with results reportedelsewhere for Triton X-100 sorption (7, 16, 18, 19, 40). Theisotherm data in Figure 2 could also be described by a two-part linear isotherm indicating strong, linear uptake of TritonX-100 at sub-cmc concentrations and essentially zero ad-ditional uptake of Triton X-100 at supra-cmc concentrations,although this approach would not be consistent with ap-proaches used in previous investigations (7, 16, 18, 19, 40).Figure 3 describes the rate-limited uptake of Triton X-100 inbatch sorption experiments. Equilibrium does not appear tobe reached for approximately 96 h. The kinetic sorption

∂Sk

∂t) k[(1 - F)

abC1 + bC

- Sk] (2)

S ) Se + Sk ) abC1 + bC

(3)

[1 + FFab

(1 + bC)2]∂C∂t

+ Fθ∂Sk

∂t) ∇‚(D∇C) - ∇‚(vC) + G

(4)

C(x,0) ) Se(x,0) ) Sk(x,0) ) 0 (5)

(C - Dv∂C∂x)|

x)0) Cin (6)

∂C(∞,t)∂x

) finite (7)

3568 9 ENVIRONMENTAL SCIENCE & TECHNOLOGY / VOL. 31, NO. 12, 1997

parameters F and k for the model fit of the data in Figure 3are given in Table 2.

Figure 4 presents the effluent concentrations of TritonX-100 from triplicate column experiments as a function oftime. The break-through curves are approximately S-shaped,although there is considerable tailing, indicative of nonequi-librium (e.g., rate-limited) sorption. Two model fits are shownfor each data set in Figure 4. The model fits indicated by thedashed lines are based on the assumption that a localequilibrium exists between the soil and water concentrationsof Triton X-100 (e.g., F ) 1 in eqs 2 and 4). The model fitsindicated by the solid lines do not assume a local sorptionequilibrium (e.g., F < 1). The average kinetic sorptionparameters (F and k) for the solid-line model fits of the datain Figure 4 are given in Table 2. The linear velocities (v) anddispersion coefficients (D) were determined by calibration ofthe model to the effluent concentrations of the nonsorbingBr- tracer (data not shown). Values of v and D for the columnexperiments are given in Table 3. The observed variabilityin the dispersion coefficient in Table 3 is likely caused bydifferences in the column packing.

Figure 5 presents the results of the Br- tracer test at thefield site. Measured and simulated tracer concentrations andmodel fits are given as a function of time for wells 92-17,

92-18, and 92-19. The longitudinal and tranverse disper-sivities were calibrated to the field tracer data to obtain themodel fit in Figure 5. Three values of longitudinal dispersivity(0.08 m, 0.24 m, and 1.04 m) were used to calibrate the modelto the observed data with the magnitude of the longitudinaldispersivity increasing with distance from the injection well.The transverse dispersivity was assigned a value equal to one-tenth the longitudinal dispersivity.

Figure 6 presents the Triton X-100 concentration data inwells 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 dataassuming that a local sorption equilibrium can be assumed.Only the model fit for well 92-17 is shown, as model fits forthe other wells do not appear until times greater than 200 d.Graph B in Figure 6 gives the model fit for the surfactant data

FIGURE 2. Equilibrium isotherm for Triton X-100 sorption from waterto soil.

FIGURE 3. Aqueous Triton X-100 concentration versus time in soilwater batch reactors.

TABLE 2. Values of Parameters F (Fraction of EquilibriumSorption Sites) and k (Mass-Transfer Coefficient) forSimulations of Triton X-100 Sorption to Soil from Water forLaboratory Batch and Column Experiments and a FieldExperimenta

parameterbatch

experimentscolumn

experimentsfield

experiment

F 0.47 0.19 0.15k (s-1) 3.1 × 10-6 1.5 × 10-6 1.5 × 10-7

a All experiments used the same soil type.

FIGURE 4. Triton X-100 effluent concentrations versus time andequilibrium and kinetic model fits for triplicate column experiments.

TABLE 3. Parameters Used To Simulate Transport of TritonX-100 through Triplicate Soil Columnsa

column v (cm/h) D (cm2/h) column v (cm/h) D (cm2/h)

A 6.15 20.9 C 6.48 45.2B 6.56 147 mean 6.40 71.2

a The average velocities (v) and dispersion coefficients (D) weredetermined from Br- tracer tests and optimized model fits.

VOL. 31, NO. 12, 1997 / ENVIRONMENTAL SCIENCE & TECHNOLOGY 9 3569

assuming that sorption is rate limited and can be describedby the two-site formulation.

DiscussionThe experimental data and analyses presented in the previoussection indicate that Triton X-100 sorption to natural soil israte-limited and cannot be accurately described by anequilibrium sorption model. This observation is true forlaboratory batch and column experiments (Figures 3 and 4)and for the field experiments (Figure 6). The rate-limitedsorption of Triton X-100 in laboratory experiments has beenreported previously (7), but the comparison of sorption ratesfor a single surfactant and soil under different experimentalconditions (e.g., batch, column, and field) provides additionalinsight 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 valuessuggests that mass-transfer rate limitations are occurring atboth the particle scale and at the porous-media scale. As thetype of experiment changes from batch to column to field,the parameter F, which equals the fraction of equilibriumsites, decreases; the parameter k, which is the mass-transfercoefficient between soil and water, also decreases. Therefore,

the sorption rate limitations become greater as the experimenttype changes from batch to column to field. In the batchsystem, the soil water slurry is being mechanically mixedduring incubation. Therefore, any rate-limited sorption iscaused at the particle scale by either diffusion into the soilorganic matter or diffusion into intraparticle pore channels.In the column experiments, heterogeneities and dead-endpores are likely introduced despite attempts to pack the soiluniformly into the soil column. Therefore, decreases in Fand k as the experimental procedure is changed from batchto column is indicative of a “dual porosity” system. Part ofthe pore space is filled with mobile water that can advect theTriton X-100 (e.g., the effective porosity) whereas anotherpart of the pore space contains stagnant water. Transfer ofthe solute between the mobile and immobile regions canonly occur by molecular diffusion. At the field scale,heterogeneities become more pronounced than those foundin the column experiment. In the field, zones of significantlyvarying permeabilities may exist in the aquifer in the regionof the surfactant transport experiment. Furthermore, chemi-cal (sorption) heterogeneities may also contribute to the strongkinetic effects observed in the field experiment (41). Chemicalheterogeneities arise from soil with varying sorptive properties(e.g., different organic carbon contents, intraparticle porosity,etc.). Although soil samples used in laboratory experimentswere collected during the drilling of the monitoring andinjection wells for the field test, it is likely that the chemicalheterogeneities encountered by the surfactant as it is trans-ported through the aquifer are greater than those encounteredby the surfactant in the soil column experiments. Thesephysical and chemical heterogeneities require further de-creases in the kinetic sorption parameters F and k to allowadequate model fits of the column and field data.

It should be noted that if physical heterogeneities areindeed contributing to the decreasing values of F and k in thecolumn and field experiments relative to the batch experi-ments, the meaning of these two parameters becomesambiguous. For example, considering only physical hetero-geneities, a more appropriate mathematical definition of theparameters would be the fraction of mobile water (F) and themass-transfer coefficient for solute transport between mobileand immobile water (k). Because both rate-limited sorptionand physical heterogeneities are affecting the transport ofTriton X-100, a more physically realistic model of the fieldsystem may have to consider three phases: soil, mobile water,and immobile water. Mass-transfer limitations would existbetween all three of these phases.

The effects of field-scale physical heterogeneities are alsoevident from the Br- tracer-test data. Although Br- is generallyconsidered to be a conservative (nonsorbing) tracer, it wasdifficult to obtain a precise match of the field tracer data withthe solute-transport model and a constant value of thedispersivity (Figure 5). Review of Figure 5 shows that thepeak Br- concentrations in wells 92-18 and 92-19 occur almostsimultaneously and all of the peaks are nonsymmetric withextended “tails”. Cross-sectional maps of the field sitedeveloped by the U.S. Geological Survey (27) show severallayers of lower permeability silty sand interspersed throughoutthe sand-and-gravel aquifer, offering further evidence ofphysical heterogeneities in the study area. To obtain themodel fits shown in Figure 5, the dispersivity had to beincreased with longitudinal distance from the injection well.This is equivalent to increasing the dispersivity for the entiremodel domain over time (e.g., as the surfactant pulse movesdowngradient, larger dispersivity values are needed to accountfor the larger range of heterogeneities it encounters) and isconsistent with the recognized observation that dispersivitiesincrease with the scale of the problem (42-45). Therefore,it is apparent that both particle scale and porous-media scalediffusion processes are contributing to the rate-limited masstransfer of Triton X-100 in the aquifer at Picatinny Arsenal.

FIGURE 5. Concentration of Br- in groundwater samples collectedfrom three wells as a function of time and model fit of the data.

FIGURE 6. Concentration of Triton X-100 in groundwater samplescollected from four wells as a function of time and equilibrium(graph A) and kinetic (graph B) model fits of the data.

3570 9 ENVIRONMENTAL SCIENCE & TECHNOLOGY / VOL. 31, NO. 12, 1997

There may be two additional aspects of the experimentaldesign that influence the tracer and surfactant transportexperiments. First, the relatively high concentration of thetracer solution injected into the aquifer may have causeddensity-dependent flow near the injection well. This wouldresult in greater “sinking” of the Br- plume than the surfactantplume because of its slightly greater density. Second,simulation of solute transport is performed with a two-dimensional, vertically averaged model. Given that themonitoring and injection wells are not screened over the entireaquifer, it is possible that there were significant verticalconcentration gradients in the aquifer during the fieldexperiment. Both these aspects are likely not problematicbecause of the expected large extent of vertical mixing in thevicinity of the injection well casing caused by the large,induced vertical hydraulic gradient relative to the horizontalgradient. Therefore, it is likely that this vertical mixing hasa much more significant effect on the tracer transport thandensity effects and strengthens the assumption of verticalaveraging for both the tracer and the surfactant. Furthermore,if signficant vertical concentration gradients exist, these arepresent for both the tracer and the surfactant, and the model-calibrated dispersivity accounts for the information loss dueto vertical averaging.

The rate limitations observed in this research at thelaboratory and field scale have many important implicationsfor surfactant-based remediation technologies. First, becauseof the observed rate-limited sorption, Triton X-100 that isinjected into the subsurface will be transported significantlyfarther in a given time period than a local-equilibrium sorptionmodel would predict. For example, the peak surfactantconcentration in well 92-17 predicted by the local-equilibriumsorption model in Figure 6 appears at a time of about 120 d,whereas the actual surfactant peak concentration is observedaround 40 d. This is probably a beneficial effect at the startof a surfactant remediation operation because it allows thesurfactant to be transported greater distances in a given timeperiod. Conversely, the rate-limited sorption may be prob-lematic at the conclusion of a surfactant remediation opera-tion if the surfactant must be removed from the subsurface.Because of the sorption rate limitations, it will require longertime periods to remove the surfactant from the aquifer thanif sorption was governed by a local equilibrium.

Compared to anionic surfactants and some other nonionicsurfactants, Triton X-100 sorbs relatively strongly to the fieldsoil (Figure 2), particularly for equilibrium aqueous concen-trations less than cmc. This strong sorption would increasematerial costs for engineered surfactant remediation schemes,since a large fraction of the surfactant will be associated withthe soil phase. Because of the flattening of the sorptionisotherm at concentrations above cmc, this may be lessproblematic at concentrations well above cmc. However,the strong sorption of Triton X-100 at sub-cmc to soil maybe the reason for its ability to increase the rate of desorptionof other sorbed organic pollutants such as trichloroetheneand carbon tetrachloride (3, 5). Interaction of the surfactantwith the soil organic matter reduces the diffusional resistancesof the soil organic matter and facilitates the outward diffusionof sorbed organic solutes into the bulk solution (3, 5).

AcknowledgmentsThis research has been supported by the National Center forEnvironmental Research and Quality Assurance (NCERQA)of the U.S. Environmental Protection Agency and the ToxicSubstances Hydrology Program of the U.S. Geological Survey.Part of the computational support for this research wasobtained from a grant from the IBM Environmental ResearchProgram. However, any opinions, findings, and conclusionsor recommendations expressed in this material are those ofthe authors and do not necessarily reflect the views of theIBM Corporation.

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Received for review April 7, 1997. Revised manuscript re-ceived August 15, 1997. Accepted August 18, 1997.X

ES970314V

X Abstract published in Advance ACS Abstracts, October 1, 1997.

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