Environ. Sci. Technol. 1996, 30, 3223-3231
Laboratory Assessment of BTEX Soil Flushing ALLISON A. MACKAY,† YU-PING CHIN,‡ JOHN K. MACFARLANE,† AND P H I L I P M . G S C H W E N D * ,† Department of Civil and Environmental Engineering, Ralph M. Parsons Laboratory, 48-415, Massachusetts Institute of Technology, Cambridge, Massachusetts 02139, and Department of Geological Sciences, The Ohio State University, Columbus, Ohio 43210
Soil cores from the unsaturated zone of a waste site contaminated with benzene, toluene, ethylbenzene, and xylenes (BTEX) were flushed with water under conditions representative of those proposed for site remediation. With a laboratory column dispersivity of 1 cm, kinetic modeling of the breakthrough curves gave desorption rate constants of 0.01-0.1 h-1. Desorption studies in batch tests gave similar rates after accounting for differences in the solid-towater ratios. The presence of nonaqueous phase liquid (NAPL) in one soil column was deduced from the flushing behavior of the xylenes, which showed increasing effluent concentration with increasing time. The amount of NAPL was estimated to be 5 mgNAPL/ gsoil from the flushing of the most water-soluble constituent, benzene. This quantity of NAPL would increase the site’s estimated flushing time by more than a factor of 5 over the case with only kinetically limited desorption of BTEX from soil particles.
Introduction The soil and associated groundwater at many sites throughout the world have become contaminated by organic compounds through spills and disposal. Where the contaminants are relatively soluble (i.e., aqueous solubility > 0.1 g/L), remediators have proposed exploiting this chemical property by using vadose zone flushing and pump-andtreat technologies. Field experience, however, has led experts to conclude that pump-and-treat is often an ineffective technique (1, 2). Levels of contaminants have been seen to persist in the flush water for long periods of time or have increased after the pumps have been turned off. Understanding which subsurface process(es) limit(s) such pump-and-treat may aid in optimizing or enhancing this remediation approach or in evaluating whether this method will even be viable. Vadose zone flushing, an extension of pump-and-treat technology, can be limited by a number of processes. First, * Corresponding author e-mail address:
[email protected]; fax: 617-258-8850. † Massachusetts Institute of Technology. ‡ The Ohio State University.
S0013-936X(95)00963-1 CCC: $12.00
1996 American Chemical Society
field-contaminated soils (3-7) have exhibited slow grainscale desorption kinetics; that is, release time scales are longer than those predicted from the behavior of laboratoryspiked compounds (3, 7). This lowers the flushing velocities that can be used to maximize mass removal rates. Secondly, the heterogeneous nature of soil suggests that regions of low hydraulic conductivity may exist from which contaminants must diffuse over long periods of time to more effectively flushed layers. Finally, flushing may be hindered by the presence of residual nonaqueous phase liquids (NAPLs), too finely dispersed to be displaced (8), and thus only removed by dissolution. Identifying which process is limiting the cleanup at a particular site may suggest approaches for enhancing cleanup. For example, measures must first address residual NAPL, if it is present (excavate key regions, steam displacement, surfactant/cosolvent flush), but such remedial activities (e.g., surfactant flush) may not be a substantial aid if contaminants are leaching slowly from clay lenses. Here, we describe a laboratory method used to assess which mechanisms (if any) limit flushing as a remediation strategy at a particular field site being evaluated for efficiency of various cleanup approaches. This Midwestern site’s soil was a sandy-loam contaminated with benzene, toluene, ethylbenzene, and xylenes (BTEX) from just below the ground surface (ca. 6 ft) to the water table (ca. 20 ft). These aromatic hydrocarbons were disposed in drums and discharged into lagoons over a period of 5 years in the late 1960s when the site was used for illegal waste disposal. The proposed remediation scheme was to apply water at the surface of the soil and flush vertically by gravity infiltration. After flushing the contaminated soil, the water would be recovered using pumping wells for treatment at the land surface. Our laboratory method attempted to mimic these remediation conditions by flushing undisturbed soil columns at site-specific groundwater velocities. This allowed comparison of the time scale for desorption or dissolution in the intact deposit with that for flushing. Subsequently, we evaluated the impacts of (1) rate-limited desorption and (2) NAPL presence on the proposed site cleanup time under the assumption that our isolated cores reflected field hydraulics.
Method Reagents. The target compounds of interest were benzene, toluene, ethylbenzene, and xylenes. Because the meta and para isomers of xylene co-eluted from the chromatographic column used, the combined total of both isomers is reported as m,p-xylene. Standards for the target compounds were prepared in methanol (spectroscopic grade, EM Science, Gibbstown, NJ) from 99+% pure compounds, used as received from Alltech Associates (Deerfield, IL). Internal standards of 1,4-bromofluorobenzene, 1,4-difluorobenzene, 1-chloro-n-pentane, and 1-chloro-n-heptane were prepared in methanol from 99+% pure compounds and used as received from Aldrich Chemical Co., Inc. (Milwaukee, WI). Other reagents used were sodium bicarbonate (Mallinckrodt, Paris, KY), sodium azide, and mercuric chloride (both Fluka Chemie, Switzerland). Sample Collection. To replicate the field conditions as closely as possible, a flushing apparatus was constructed
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FIGURE 1. Apparatus for flushing intact soil cores.
that could accept soil cores collected at the field site (Figure 1). Soil samples were collected using 5.1 cm diameter by 76 cm long, thin-walled, stainless steel Shelby tubes. The coring profile spanned a depth of 1.8-6 m (6 ft to 20 ft) below the ground surface. These soil samples were taken consecutively by sampling ahead of an auger, drilled to depth. The Shelby tubes were immediately capped in the field and stored at 4 °C for transport back to the laboratory. Damaged ends of the Shelby tubes with some core material were removed at the laboratory to facilitate fitting the tubes into the end caps of the flushing apparatus. All apparatus fittings and tubing were aluminum, stainless steel, or glass to minimize sorptive losses of BTEX. The columns were flushed with reverse osmosis (RO) water (Bridgewater, PA) containing 5 mM sodium bicarbonate to buffer the pH at 8.3 and 1.5 mM sodium azide to minimize microbial activity. We observed very little fines flushed from the column, despite the sodium solution chemistry. A constant head of 1 m was maintained by Mariotte bottles. Darcy’s law was used to calculate hydraulic conductivities based on timed collection of a fixed volume of water. Observed hydraulic conductivities ranged from 10-4 to 10-3 cm/s. These are typical values for saturated silty sand (9), supporting our subsequent assumption of saturation of the soil columns. Column flushate was collected in 100-mL glass syringes connected to the column outlet by Luer-Lok fittings, eliminating headspace over the samples. Syringes were replaced and analyzed approximately daily. The flushing systems were maintained at 23 ( 3 °C throughout the experiment. The flushing apparatus was designed to mimic the field remediation flow conditions by installation of flow control valves at the base of the columns (Figure 1). In the field, each soil segment is not isolated hydraulically; the rate of flow is influenced by the soil located above and below. Therefore, the flow of water through each column was regulated at its base, simulating the resistance provided by the remainder of the soil column, as in the field. The valves
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were adjusted to give flushate flow rates through the column of about 5 cm/d (Table 1). Another core was also taken a few feet from the Shelby tube samples using a split spoon sampler with 15 cm brass inserts. This provided soil, representative of the horizons which the Shelby tubes spanned, for analysis of physical properties. Organic carbon content and porosity were analyzed in these samples by Versar (Springfield, VA) (Table 1). Batch Desorption Kinetics. Desorption kinetics from the field-contaminated soil were also examined in batch mode. The soil used to observe desorption kinetics was from a column flushed for 16 pore volumes. At this time, benzene and toluene were no longer detectable in the effluent. Ethylbenzene and xylene concentrations were 10% of their initial values. The soil was stored at 4 °C for 4 months after its use in the flushing apparatus before use in batch studies. Soil (5.0 g) was placed in a 50-mL (nominal volume) centrifuge tube (with Teflon-lined screw cap) with enough water (∼52 mL) to fill the tube leaving no headspace. The water contained 5 mM sodium bicarbonate, 1.5 mM sodium azide, and 40 µM mercuric chloride. The mercuric chloride was added because a previous experiment had shown 1.5 mM sodium azide to be ineffective in the batch experiments after a few days. Centrifuge tubes were tumbled (6 rpm) in the dark at 30 ( 1 °C. At each time point, three tubes were removed from the tumbling apparatus. They were centrifuged at 1200g for 30 min to separate the supernatant and the soil. Analysis. Flushate and supernatant samples were chiefly analyzed by capillary gas chromatography with flame ionization detection (GC-FID). Internal standards of 1,4bromofluorobenzene, 1,4-difluorobenzene, 1-chloro-npentane, and 1-chloro-n-heptane were used in all analyses to assess recoveries. These standards were added to samples to be analyzed by direct aqueous injections at concentrations of 50 mg/L; samples to be analyzed via purge-and-trap procedures were spiked at 50 µg/L. Samples with target compound concentrations in the microgram per liter (ppb) range were concentrated with a Tekmar LSC2000 purge-and-trap concentrator. Aqueous samples (5 mL) were purged for 7 min with helium at 10 mL/min. Volatiles in the purge gases were concentrated on a Tenax/ silica gel/charcoal trap. The trap was desorbed for 4 min at 180 °C at a flow rate of 20 mL/min. The desorbed sample was transferred directly to the head of the GC capillary column through a 0.32 mm i.d. deactivated fused silica line held at 150 °C. The GC employed was a Carlo Erba HRGC 5160 equipped with a 60 m, 1 µm film thickness, DB5 capillary column (5% phenylmethylpolysiloxane, J&W Scientific, Folsom, CA). The temperature program began with 3 min at 35 °C, followed by a ramp at 10 °C/min to 125 °C, a 20 °C/min ramp to 200 °C, and a 15 °C/min ramp to 225 °C, where the temperature was held for 5 min. Helium was used as the carrier gas at 20 mL/min. The compounds were detected by a flame ionization detector with a base temperature of 250 °C. Between samples, the purge trap was reconditioned by baking at 225 °C for 5-15 min, depending upon the extent of sample contamination. Blanks were run to verify no carryover. Water samples with mg/L (ppm) concentrations were analyzed by direct aqueous injection to a Carlo Erba Fractovap, equipped with a 30 m, 0.3 µm film thickness, RTX-1 capillary column (100% phenylmethylpolysiloxane, Restek Corp., Bellefonte, PA) and FID. Cold on-column
TABLE 1
Physical Soil Properties and Experimental Column Flushing Conditionsa initial effluent concn (µg/L) depth (ft)
porosity
av foc
flushing column length (cm)
6-8.5 8.5-11 11-14 14-17 17-20
0.40 0.39 0.26 0.40 0.36
0.0022 0.0011 0.0036 0.0012 0.002
38 45 58 53 30
a
av flushing velocity (cm/h)
benzene
toluene
ethyl benzene
o-xylene
0.13 0.35 0.37 0.3 0.37 solubility
380 110 140 50 5 1 780 000
36 000 12 000 12 000 7 700 2 500 580 000
10 000 4 200 3 600 2 400 1 600 190 000
12 000 5 600 4 600 3 300 2 100 220 000
Solubility data reported at 25 °C from Miller et al. (24).
injection (1-µL aliquot) was utilized with a temperature program beginning at 100 °C with a 10 °C/min ramp to 200 °C. Helium carrier flow was at 2 mL/min. Blanks were run to verify no carryover between runs. Identities and purities of analyte compounds in the flushate solutions were initially confirmed by comparison of mass spectra and retention times with those of known standards. Compounds were quantified by measuring peak heights, corrected for internal standard recoveries. Peak heights were compared to those generated using injections of known quantities of external standards. A five-point calibration curve was developed for each target compound over the range of measured concentrations to verify linear detector response. Error in analyses was estimated from the variation in internal standards, which showed a relative standard deviation of (13% from 500 analyses over 3 months. Soil analysis was performed by Versar (Springfield, VA). Fraction organic carbon (foc) was determined using a Perkin Elmer 2400 elemental analyzer. Soil porosity was calculated from bulk (10) and particle density (11). EPA Method 8260 (12) was used for analysis of (1) initial soil concentrations of target compounds in the duplicate core and (2) final soil concentrations after flushing the soil columns. While Method 8260 may yield measures of soil concentrations that are lower than the reality, we only used these results for comparative purposes (e.g., to identify heavily contaminated horizons, to evaluate the residual contamination in flushed cores). Modeling Column Effluent Profiles. The soil column effluent concentrations were modeled with the onedimensional advection-dispersion equation using CXTFIT (13). Few observations of contaminant removal from fieldcontaminated sites have been made using column studies (4, 7). In our study, long columns and relatively slow porewater velocities were employed, and thus equilibrium sorption conditions may have been attained. The likelihood of sorptive equilibrium under these flow conditions was assessed using the local equilibrium assumption (LEA) in the model. With this assumption, CXTFIT was used to fit the effluent profiles by adjusting two parameters: the dispersion coefficient (D) and the retardation factor (R). Column partition coefficients were then calculated from R ) 1 + rswKd, where rsw (g/mL) is the solid-to-water ratio and Kd (mL/g) is the solid-water partition coefficient. Fits for kinetically limited column flushing were also obtained with CXTFIT. This modeling program employs a two-site model in which a fraction, (F), of all sorption sites still adheres to the LEA case (13). To limit fitting parameters because F cannot be determined experimentally, a single mass transfer
rate constant was obtained by setting F ) 0 [i.e., CXTFIT parameter β ) 1/R (13, eq 37)]. With a fixed dispersivity of 1 cm (discussed in Results), the fitted parameters for the kinetic case were R values and the desorption rate constants were ω (h-1). Optimal fits were found by using a value of 1/β updated to equal the previous best fit R. Optimal fits were recognized when the best fit R equaled the input choice of 1/β. The soil columns recovered from the field were initially unsaturated when flow was applied; however, the initial conditions of the CXTFIT model assume that the column is saturated with porewater in equilibrium with the contaminated soil phase. That is, when flow is initiated to the column (t ) 0+), the effluent concentration would be the initial equilibrated porewater concentration. Obviously, a length of time equal to the passage of one pore volume minus the residual water content was required to saturate our columns with water. (Note we calculate that the fraction of BTEX compounds displaced in soil gas would be e6% in all cases.) For a sandy loam soil, the residual water is typically 10-15% or about one-third to one-half the porosity (14, 15). Therefore, the experimental time frame for effluent collection (t ) 0 when flow from base of column) deviated by 2/3 pore volume from the modeling time frame of CXTFIT (tCXT ) 0 when flow initiated to top of column). To reconcile the two for modeling purposes, the time necessary to displace 2/3 of a pore volume was added to each experimental time point to offset it into the CXTFIT time frame, such that the first observed effluent now occurs at tCXT ) 2/3 pore volume for modeling. The initial effluent concentration, Ci, in equilibrium with soil at tCXT ) 0 must be set to model effluent data. This required us to construct an imaginary breakthrough curve, and set a theoretical initial effluent concentration, (Ci*) for the first 2/3 pore volume so that CXTFIT could be used to model our experimental data. The maximum value for the Ci* was estimated by assuming that the theoretical first 2/3 pore volume was fully equilibrated with the soil when it left the column. The subsequent soil concentration at tCXT corresponding to 2/3 pore volume would then be (1 - fw) × Ci* where fw is the fraction of total contaminant mass in the initial flush water:
fw )
1 1 + rswKd
(1)
For an average column porosity of 0.36 and partition coefficients ranging from 0.4 mL/g for benzene to 2 mL/g for o-xylene, respectively, the mass removed from the column by the first 2/3 pore volume would be about 30%
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a
b FIGURE 3. Effluent profiles of BTEX in column 8.5-11. Concentrations were normalized to the first observed effluent concentration to allow comparison of flushing trends for compounds present at different levels. Representative error bars are shown only for selected points to improve figure clarity. Errors were calculated from the 13% standard deviation of 500 internal standard measurements.
where C (µg/mL) is the dissolved concentration and S (µg/ g) is the sorbed concentration. The batch desorption rate constant, k (h-1), was determined by nonlinear least squares fitting (SigmaPlot, Jandel Scientific) of batch data to the solution to eq 2 for a finite bath (16):
C ) C∞[1 - exp(-k(Kdrrw + 1)t)]
(3)
where C∞ is the equilibrium aqueous concentration. FIGURE 2. Sensitivity of kinetic fit to choice of Ci* for (a) benzene and (b) o-xylene in column 8.5-11. Least squares fitting parameters are detailed in text.
for benzene and 10% for o-xylene. Thus, Ci* should be about 1.4Ci for benzene and 1.1Ci for o-xylene. The effect of our imprecise knowledge of Ci* on the subsequently fit parameters (R and ω) was assessed with the kinetic model of CXTFIT. The fitted breakthrough curves were certainly sensitive to the choice of Ci* (Figure 2); however, the sum of squares and correlation coefficient values for the fits were the same for each Ci* value. To exemplify the magnitude of effects on the fitting parameters, we give the values of R and ω ((95% confidence limits) for benzene and o-xylene flushing from column 8-11. If Ci* ) 1.3Ci, then R ) 2.2 ( 0.1, ω ) 0.014 ( 0.003 h-1; if Ci* ) 1.4Ci, then R ) 2.1 ( 0.1, ω ) 0.011 ( 0.002 h-1; if Ci* ) 1.5Ci, then R ) 2.1 ( 0.1, ω ) 0.010 ( 0.002 h-1. For o-xylene, if Ci* ) 1.0Ci, then R ) 9.8 ( 0.2, ω ) 0.008 ( 0.002 h-1; if Ci* ) 1.1Ci, then R ) 9.4 ( 0.2, ω ) 0.006 ( 0.001 h-1; if Ci* ) 1.2Ci, then R ) 9.1 ( 0.3, ω ) 0.004 ( 0.001 h-1. Hence, the best fit values of R and ω have imprecisions of about 5 and 15%, respectively, resulting from the uncertainty in the choice of Ci* . Modeling Batch Desorption Kinetics. The increase in batch supernatant concentration was modeled as a firstorder mass transfer process:
∂C ) krsw (S - KdC) ∂t 3226
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(2)
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Results and Discussion Flushing Behavior and Equilibrium Modeling. The initial effluent concentrations were lower for deeper soil samples (Table 1). This suggests that contaminants leached down from the ground surface. For most of the field cores flushed in the laboratory, there was a marked decline in compound effluent concentrations with time (Figure 3). As expected, compound concentrations declined according to solubility: the more water-soluble compounds (e.g., benzene) being flushed out more quickly. In column 8.5-11 effluent benzene concentration was reduced from 100 µg/L to 0.5 (i.e., 200 times) in 5 pore volumes. In 10 pore volumes, toluene effluent concentration declined from 12 300 to 400 µg/L (i.e., 30 times). After flushing for 15 pore volumes, the effluent concentrations of ethylbenzene and xylenes were each only reduced by 1 order of magnitude to about 1000 µg/L. The CXTFIT/LEA model was first fit to the effluent curves to determine if the breakthrough curves were consistent with equilibrium sorption conditions at the low porewater velocities employed. Partition coefficients calculated from best fit R values were compared to predicted values (Kd ) focKoc). Linear isotherms have been observed for other soils and aquifer solids with similar organic carbon contents (17, 18). Our observed Kd values agreed well with predicted Kd for columns 14-17 and 17-20 (Table 2) but deviated substantially for the shallower cores. For example, observed Kd values for column 8.5-11 were about three times greater than the corresponding estimated focKoc products, while
TABLE 2
Partition Coefficients (mL/g) Determined from LEA Fitting of Column Effluent Profilesa column
benzene
toluene
ethyl-benzene
m,p-xylene
o-xylene
8.5-11
0.27 0.12 0.25 0.38 0.07 0.13 0.42 0.21 2.02
0.78 0.29 0.36 0.95 0.20 0.32 0.52 0.53 2.40
2.3 0.63 0.85 2.1 0.43 0.69 0.88 1.2 2.74
2.5 0.66 0.61 2.2 0.48 0.72 0.91 1.2 2.79
2.1 0.60 0.32 2.0 0.42 0.65 0.85 1.1 2.74
11-14 14-17 17-20 log Koc aItalicized
values were calculated from Kd ) Kocfoc with Koc from Schwarzenbach and Westall (17). 95% confidence level was (5% of Kd value.
TABLE 3
Column Dispersivity (cm) Values Determined from LEA Fitting of Effluent Profilesa
a
column
benzene
toluene
ethylbenzene
m,p-xylene
o-xylene
8.5-11 11-14 14-17 17-20
10 4 2**
8 5 2**
7 6* 6 7
7 7 7
9 8 7 7
95% confidence level was (15% of dispersivity, with these exceptions: (*) (50%, (**) (60%. Dashes (-) denote fitting errors greater than 100%.
TABLE 4
Column Desorption Half-Lives (h) Determined by Kinetic Fitting of the Effluent Profiles, Assuming r ) 1 cma
a
column
benzene
toluene
ethyl-benzene
m,p-xylene
o-xylene
8.5-11 11-14 14-17 17-20
90 50 -
100 60 10**
130 110* 60 90
160 70 90
170 78 70 90
95% confidence level was (15% of half life, with these exceptions: (*) (50%, (**) (75%. Dashes (-) denote fitting errors greater than 100%.
observed Kd values for the 11-14 section were about three times less than calculated. Such deviations may have resulted from inaccurate foc characterization arising from the use of a replicate core, not representative of the flushed column. In general, the foc values varied by a factor of 4 from 0.001 to 0.004, indicating the accuracy of our estimated Kd ()focKoc) should be no better than our knowledge of the foc. We conclude that these data alone are not sufficient to preclude the hypothesis that the columns exhibited sorptive equilibrium. The column dispersivities, R, were calculated from the best fit dispersion coefficients using D ≈ RU where U (cm/ h) is the flushing velocity (Table 3). The deduced R values ranged from 4 to 10 cm, generally greater than previously reported for laboratory columns. For flow through disturbed and undisturbed, unconsolidated soils, R ranges from 0.1 to 2 cm (9). This length scale is generally thought to correspond to the effective grain sizes in the soil columns being tested, and we know that several-centimeter grains in our cores did not exist (as evidenced by X-rays of two intact cores and by observation of the core contents after flushing). Consequently, we believe these fitted R values must reflect additional factors, such as sorptive disequilibria not included in the LEA model. If such flushing behavior were fit incorrectly, that is assuming sorptive equilibrium, the resulting dispersivity would be substantially greater than that obtained with a conservative tracer. Kinetically Limited Desorption. As the large dispersivities obtained with LEA modeling did not agree with
expected values nor physical observations, a kinetic interpretation of sorption was introduced. Choosing a reasonable column dispersivity of 1 cm, we modeled the “excess” dispersivity as kinetically limited flushing. (Due to the contractual arrangements allowing us access to these cores, we could not run a conservative tracer test to assess column dispersivity directly.) The retardation values and hence Kd values, obtained with these fits did not deviate within the fitting error from those obtained with LEA modeling. Best fit desorption rates for the BTEX indicated characteristic desorption times (i.e., ω-1) were about 100 h (Table 4). Setting R to 0.1 cm as a reasonable lower limit, best fit retardation values are again similar to the LEA case, but the best fit desorption times would have been on the order of weeks [e.g., column 8.5-11: ω-1 (benzene) ) 24 d, ω-1 (o-xylene) ) 62 d]. Our observed decline in soil concentrations is consistent with R near 1 cm, rather than 0.1 cm. While duplicate cores may not be representative, comparison of initial soil concentrations in a replicate core with concentrations in the soil columns after flushing did suggest that a large fraction of the BTEX mass was lost. For example, the o-xylene soil concentration in column 8-11 was only 1% of its concentration in the replicate core material after flushing for 60 days. Assuming a first-order decrease in soil concentration, such mass removal would have a characteristic time of 300 h. This is comparable to the characteristic time of 170 h obtained if R ) 1 cm, but not to 60 days if R ) 0.1 cm. Moreover, the integrated BTEX
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function of the product, rswKd (19). Thus, even for the same sorbate, systems with greater rsw (e.g., our columns vs the batch tests) will exhibit a faster approach to equilibrium. The exact solution has been approximated with a single mass transfer coefficient (16, 20). Assuming a diffusion desorption mechanism and the same diffusion distances in both batch and column cases, an expression relating the effect of rsw on kinetics would be given by (16, eq 41):
k(rsw1) k(rsw2)
)
11rsw1Kd + 23 11rsw2Kd + 23
(4)
For the rsw in the columns (4.25 g/mL) vs that in the batch (0.09 g/mL), the desorption rate constants for ethylbenzene and the xylenes should be about four times faster in the column. Therefore, batch characteristic times of 300 h would correspond to about 70 h in the columns. This observed time supports the conclusion that the flushing of the soil columns reflected slow desorption kinetics, even at the low flow rates that we employed. Thus, under other ambient groundwater conditions, and even more so under induced pumping conditions, such desorption kinetics would cause longer times than predicted assuming sorption equilibrium to flush the soil to the same cleanup levels.
FIGURE 4. Batch desorption kinetics for ethylbenzene and o-xylene.
masses removed in the flushing effluent after about 2 months (i.e., several times a characteristic desorption time of 2-7 days) were very large compared to the measured residual BTEX masses left in the same columns after flushing was terminated. This would not have been the case if the true desorption time scale had been of the order of 3-7 weeks. Thus, we conclude that the column dispersivities were about 1 cm and a kinetic description of column flushing is necessary. Batch Studies. To augment the kinetic interpretations of the column flushing data, we also measured desorption rates in batch experiments. Although this soil had been partially flushed, it sat undisturbed for more than 4 months before batch experiments were undertaken. Thus, we believe these batch tests should reflect desorption from a pool of sorbates distributed throughout the solid aggregates much as was the case for the columns, rather than removal of a highly resistant fraction remaining after leaching. Keeping in mind the possibility of shortening the diffusionlimiting length scales of interest, batch testing of the previously flushed solids should therefore reflect the same or faster mass transfer kinetics as seen in the columns. The ethylbenzene and xylenes desorbed over periods of hundreds of hours (Figure 4). The fitted rate constants correspond to characteristic times (k-1) of 300 ( 90 h for ethylbenzene and the xylenes. To allow comparison with the column rate constants, ω, these batch desorption rates must be adjusted for the difference in solid-to-water ratios in the batch and column. In cases of desorption into limited closed volumes, the rate of change in average sorbed concentration in time is a
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Mechanistic Interpretation. Differences in the column desorption rates for the BTEX compounds support the hypothesis that retarded radial diffusion causes the slow desorption in these soils. Characteristic desorption times were a strong function of Kd for all combinations of sorbates and sorbents (Figure 5). This type of relationship would be expected for retarded diffusion through aggregated media. An alternative rate-limiting mechanism involves diffusion into organic films (21). Such film diffusion is expected to be strongly dependent on sorbate molecular weight (reflecting the sorbate size influence on diffusivity.) Our data indicated some dependency on sorbate molecular weight, but the characteristic desorption time scales in all our cores did not “collapse” to the same trend as we saw for correlation with Kd. This could be due to changing organic film properties in these cores. However, the variations in column desorption times among sorbate/ sorbent combinations appear more consistent with retarded diffusion through soil aggregates for our particular cases. Nonaqueous Phase Liquids. One soil column exhibited markedly higher initial effluent concentrations (Table 1) and different flushing profiles (Figure 6). Nonaqueous phase liquid (NAPL) presence was indicated by initial effluent concentrations at significant fractions of solubilities: benzene, 0.37 mg/L (0.02% solubility); toluene, 36 mg/L (6%); ethylbenzene, 9.9 mg/L (6%); m,p-xylene, 34 mg/L (20%); and o-xylene, 12 mg/L (7%). Though not target compounds, isomeric C3-benzenes were each present at 3-5% of their solubilities. These data would be consistent with a NAPL composed of about 50% BTEX plus C3benzenes and the remainder composed of less soluble hydrocarbon constituents (assuming ideal solution behavior). The second line of evidence for NAPL presence in this intact soil core involved the differential flushing of the target compounds. The solubility of a compound dissolving from NAPL is proportional to its mole fraction in the nonaqueous phase (22). The mole fractions of the more hydrophobic compounds will increase as the most water-soluble com-
becomes
R ) 1 + rswKd ) 1 + rswfNAPLKNAPL
FIGURE 5. Variation of characteristic column desorption time with partition coefficient and molecular weight. The correlation coefficient of desorption time with partition coefficient for all columns was 0.84. The correlation coefficient of the log of desorption time with log of molecular weight for all columns was 0.59.
pounds are leached from the NAPL; over time this effect results in increasing flushate concentrations of the more hydrophobic compounds. Such behavior was observed in the effluent profiles of column 6-8.5 (Figure 6). Benzene, the most soluble compound, declined about 1 order of magnitude over 10 pore volumes. Effluent toluene concentrations showed minimal decline after flushing the column with 10 pore volumes. The least water soluble compounds, ethylbenzene and o-xylene, showed increasing effluent concentrations with continued flushing (Figure 6c,d). These trends in effluent concentrations also support the hypothesis that the BTEX were dissolving from a NAPL. Subsequently, when we dismantled the column, microscopic droplets of a brown immiscible liquid were observed in the porewater of this column, thereby demonstrating the presence of NAPL. The amount of NAPL initially present in column 6-8.5 was estimated from the flushing profile of the most readily leached compound, benzene. Benzene’s flushing was assumed to occur by equilibrium partitioning between the aqueous flushate and the residual NAPL. Natural organic matter sorption would have only accounted for a small portion of the observed retardation (17% assuming foc ) 0.0022) (23). The NAPL-dominated retardation factor thus
(5)
where fNAPL (mLNAPL/gsoil) is the fraction NAPL in the soil and KNAPL (mLwater/mLNAPL) is the NAPL-water partition coefficient. Since the NAPL appeared to be predominantly hydrocarbon in nature, we approximated KNAPL with the hexane-water partition coefficient, Khex (mLwater /mLhexane) from the literature (22). Fitting the effluent profile to the LEA case with CXTFIT yielded a retardation factor for benzene of 5.5. For our known rsw ) 4 g/mL and Khex for benzene of 170 mLwater /mLhexane (22), the fraction of NAPL in the soil was calculated to be 0.006 mL/g or about 5 mg/g (assuming a NAPL density of 0.85 g/mL.) Even small amounts of residual NAPL (
