Environ. Sci. Technol. 1991, 25, 274-279
(26) (27)
(28) (29)
Proposed Rule; U.S. Environmental Protection Agency. Fed. Regist. 1987,52 (No. 87), pp 16982-17050 (May 6). Charlesworth, D. H.; Marshall, W. R. AIChE J. 1960,6, 9. Goodier, J. N. Philos. Mag. 1936, 22, 678. Raask, E. Mineral Impurities in Coal Combustion; Hemisphere: New York, 1985; p 132. Bird, R. B.; Stewart, W. E.; Lightfoot, E. N. Transport
Phenomena; John Wiley & Sons: New York, 1960; pp 648-67. Received for review May 17, 1990. Revised manuscript received September 6,1990. Accepted September 17,1990. The research reported was performed under EER Corporation Subcontract 8560-49, under EPA Prime Contract 68-02-4247.
Desorptive Behavior of Trichloroethylene in Contaminated Soil Spyros G. Pavlostathis” and Kendrlck Jaglal W. J. Rowley Laboratories, Department of Civil and Environmental Engineering, Clarkson University, Potsdam, New York 13699
An investigation of the desorptive behavior of trichloroethylene (TCE) in a long-contaminated soil (silty clay) was conducted. Batch desorption of TCE reached steady state within 24 h. The temporal dependence of TCE desorption was reinforced in reequilibration studies, where a persistent fraction of TCE desorbed very slowly. The effect of pH was negligible and TCE desorption decreased with increases in ionic strength beyond 0.1 M. Batch adsorption of TCE reached an apparent steady state in 3 days. The soil adsorption partition coefficient for TCE was estimated as 11.7 mL/g. Adsorption and desorption were not reversible over short equilibration times (on the order of days). With a short soil column, continuous TCE desorption showed an initial fast rate and a subsequent slower rate, revealing the high persistence of ppb levels of TCE in long-contaminated soil.
Introduction Halogenated organic compounds, especially low molecular weight hydrocarbons, are found extensively in the environment. Some of these compounds, which are known carcinogens, have been found in subsurface soils and groundwater. Usually, since the contaminants are strongly sorbed onto the soil matrix, remediation by conventional aboveground treatment techniques is both inefficient and slow ( 1 ) . In addition, a portion of the sorbed compounds may be irreversibly bound. For example, the soil fumigant 1,2-dibromoethane was found to persist in agricultural topsoils for up to 19 years, despite its high volatility and degradability ( 2 ) . Organic molecules buried within micropores of soil aggregates can be inaccessible to microorganisms. Soil bacteria range in size from 0.5 to 0.8 pm (3)and more than half the pore volume in a silt loam soil may be represented by pores of radii less than 1 pm ( 4 ) . Consequently, the solution-phase organics should be more available to bacteria than the sorbed fraction. It would seem, therefore, that the contaminant in solution constitutes the bioavailable fraction. As biodegradation proceeds, solution concentrations are reduced and a gradient develops whereby sorbed material moves into solution. Thus, the rate of desorption of compounds from the sorbed to the aqueous phase will greatly influence the overall mineralization rate of sorbed compounds ( I ) . Studies of pollutant desorption from contaminated soils are scarce and inconclusive (51, Desorption experiments are usually performed in conjunction with adsorption studies to determine the reversibility of freshly adsorbed compounds onto the soil matrix. However, the time frame and conditions usually encountered in the field are difficult to simulate in the laboratory. 274
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A two-phase pattern of desorption of organic chemicals from soils has been revealed by kinetic studies. An initial fast stage (in hours) and a subsequent longer, slow phase (in days) have been observed. The second stage may be attributed to slow solute diffusion from internal sorption sites (6). The biphasic desorption pattern was more pronounced in soils that were contaminated for longer periods (7,8). The kinetics of desorption of a sorbed compound appears to be related to its residence time or “age” in the soil. Conditions of dynamic equilibrium and complete reversibility, usually assumed in theoretical adsorption/ desorption models, are dubious in the case of long-term contamination with hydrophobic compounds (5). A study was undertaken to investigate the desorptive behavior of trichloroethylene (TCE) from a soil that has been contaminated for at least 18 years and results are reported in this paper. Batch experiments were conducted to ascertain the influence of equilibration time, ionic strength, and pH on desorption. Continuous desorption of TCE using a soil column was also performed to closer simulate field conditions.
Materials and Methods Soil. Soil samples were obtained from a hazardous waste disposal site in New York State. The soil samples were extracted with a hollow-stem auger and split-spoon sampler and were transferred into amber glass, wide-mouth bottles, which were closed with Teflon-lined caps. Soil used in the sorption study was taken from a depth of between 5 and 6 m. When these soil samples were tested for organic contaminants (both volatile and extractable organic compounds), TCE was found to be the major contaminant. For the adsorption experiments, soil samples were oven-dried at 103.5 OC for 24 h, broken apart in a mortar with a rubber-tipped pestle, and passed through a 0.075-mm sieve. When portions of these soil samples were methanol extracted and analyzed by gas chromatography they were found to be TCE-free. Chemicals. Standards and working solutions were prepared by dissolving aliquots of neat TCE (99+%; Aldrich Chemical Co., Milwaukee, WI) in methanol (HPLC grade; Fisher Scientific, Rochester, NY) and then diluted with organic-free distilled and deionized water. Working aqueous TCE solutions without any methanol were prepared by dissolving neat TCE in organic-free, distilled, and deionized water in serum bottles with zero headspace. All working solutions contained 200 mg/L sodium azide (Sigma Chemical Co., St. Louis, MO) as a biocide. Analyses. Aqueous TCE concentrations were determined by using a purge-and-trap concentrator (Environchem, Unacon Series 810), a gas chromatography (GC) unit
0013-936X/91/0925-0274$02.50/0
0 1991 American
Chemical Society
(Varian 3700) equipped with a flame ionization detector, and a Hall detector (electrolytic conductivity detector, Tracor 1000). A 30-m, 0.53-mm-i.d. fused-silica phenyl/ methylpolysiloxane phase capillary column was used (Rtx-volatiles, Restek Corp., Bellefonte, PA). The sample was purged with nitrogen gas at a flow rate of 40 mL/min for 8 min, transferred to two traps, and again transferred to the GC column a t a flow rate of 10 mL/min by using nitrogen gas. At the exit of the capillary column, the sample was split (ca. 50/50) to the two detectors. The GC column was kept at an initial temperature of 40 "C for 5 min, then ramped a t a rate of 10 "C/min to 220 "C, and held for 5 min. Spectra-Physics 4290 integrators were connected to each of the two detectors. An Orion Research Model 701 digital pH meter equipped with a combination electrode was used for pH measurements. Total soil organic carbon was determined by the wet combustion method (9) and soil water content was measured gravimetrically. The grain size distribution of the soil was determined by sieve and hydrometer analyses for particle sizes greater and less than 75 pm, respectively (10). The soil specific surface area was determined by the ethylene glycol monoethyl ether (EGME) method (11). Soil specific gravity determinations were made in accordance with ASTM method D854 (10) and the soil cationexchange capacity was determined by following procedures outlined by Rhoades (12). Batch Studies. In the batch adsorption and desorption studies, 0.5-g (dry weight) soil samples were placed in 8-mL amber vials, which were completely filled (to minimize headspace) with the appropriate solution (aqueous TCE for adsorption and organic-free distilled water for desorption) and capped with Teflon-faced silicone septa. All experiments were conducted at a constant temperature of 20 "C and the samples were agitated continuously with a wrist-action shaker. At the end of the shaking period, the vials were centrifuged at 3000 rpm (1240g) for 30 min, and the supernatant was removed and stored at 4 "C in completely filled 4-mL amber vials for subsequent analysis. For the batch desorption kinetics, duplicate samples were prepared as described above. Supernatant TCE was quantified at 10 time intervals (0-34 h). Another kinetic desorption experiment was performed over 7 days, during which readings were taken at daily intervals. Two sets of controls were prepared and manipulated in the same manner as the samples during the desorption experiments. The first set contained the desorbing solution (i.e., organic-free distilled water) and was used as a check for contamination by TCE. The second set of controls contained a known concentration of aqueous TCE and was used as a check against TCE losses. All controls contained 200 mg/L sodium azide. In order to determine the equilibration time for the adsorption equilibrium study, adsorption kinetics were studied in batch mode. The adsorbing solution contained 100 ng/mL TCE and quantification of the aqueous TCE concentration was performed a t six time intervals (0-12 days). The time required for TCE adsorption was found to be 3 days. An equilibrium adsorption experiment was also performed. Five sets of duplicate samples containing 0.5 g (dry weight) of TCE-free soil were equilibrated for 3 days with aqueous TCE concentrations ranging from 7.14 to 50.52 ng/mL. The amount of TCE adsorbed onto the soil was calculated based on the difference of the aqueous TCE concentration at the beginning and end of the equilibration period. Preliminary adsorption tests were performed where, after centrifugation and withdrawal of the supernatant, the TCE that had adsorbed onto the soil
was recovered by two methanol extractions of the soil pellet as described in a subsequent section. On the average, 90% of the sorbed TCE-measured by the solution concentration difference-was recovered by the first methanol extraction. The second methanol extraction did not yield any significant additional amount of TCE. Two sets of controls were used in the adsorption experiment containing organic-free distilled water with 0.5 g of soil and without any soil, respectively. Sodium azide was used as a biocide at a concentration of 200 mg/L. All controls used in this study showed less than 3% TCE loss. Two successive desorption experiments with reequilibration a t every 1- or 24-h time interval were performed. The contaminated soil (0.5-g dry weight) was allowed to desorb in organic-free, distilled water for 1 h or 1day and was then centrifuged, and the supernatant TCE concentration was measured. The supernatant was then discarded and replaced with new desorbing solution, and the reequilibration process was repeated. Before the addition of the new solution, the TCE remaining in the interstitial water of the soil pellet was quantified as the product of the supernatant TCE concentration previously measured and the amount of the interstitial water, determined gravimetrically. This value was subtracted from the subsequent measured TCE to determine the actual amount of desorbed TCE during each reequilibration. Seven reequilibrations were performed for each set of both hourly and daily experiments. Batch experiments were performed to elucidate the effects of ionic strength and pH. The ionic strength was adjusted with sodium chloride to values ranging from 0.003 to 1.0 M. Additionally, the effect of the dispersant sodium hexametaphosphate was studied at doses of 0,0.4, 5, and 40 g/L. Adjustments in pH were made by the addition of either HC1 or NaOH to provide pH values of 3 , 5 , 7 , 9 , 11, and 13. The ionic strength of the pH solution was fixed at 0.01 M with sodium azide. In these experiments equilibration took place over a period of 3 days. Continuous Desorption. A continuous desorption study was performed with a short column. Details of the short column have been presented elsewhere (13). The reactor volume was 0.49 cm3 (0.37 cm long and 1.30 cm in diameter). The amount of contaminated soil used was 0.53 g (dry weight) and was taken from a depth of 5-6 m. This soil sample contained more clay than the composite soil sample used for the soil grain size analysis (see results on soil analysis). The initial soil TCE content (1981 ng/g) was determined by methanol extraction of a similar soil sample, using the procedure described in the following section. The calculated soil porosity in the short column was ca. 60%. The soil was secured in the short column with glass frits (porosity 25-50 pm; Ace Glass, Vineland, NJ) and two membrane filters (0.22- and 0.45-pm pore size; Millipore Corp., Bedford, MA) on each end. Organic-free distilled water containing 200 mg/L sodium azide was pumped through the soil column a t an average flow rate of 1.35 mL/min with a Masterflex pump (Cole Parmer Instrument Co., Chicago, IL). The eluent reservoir was a 10-L glass serum bottle with appropriate fittings to provide a nitrogen atmosphere with nitrogen replacing the used eluent. The column effluent was collected in 5-mL, Luer lock tip glass, gas-tight syringes and the TCE concentration was quantified a t periodic intervals. Extraction of TCE. The efficiency of TCE extraction from soil samples was studied by equilibrating mixtures of contaminated soil and water and soil and methanol and quantifying the TCE in the resulting solutions. Two soil TCE extraction procedures were tested: (a) a soil/ Environ. Sci. Technol., Vol. 25, No. 2, 1991 275
Table I. Results of Soil Analysisn
(I
n
U
50 I
1
W
parameter
value
sand, % silt, % clay, % specific surface, m2/g specific gravity water content (dry weight basis), % total organic carbon, % cation-exchange capacity, mequiv/100 g
56 32 12 300 2.8 22 0.13 70
E bk U
40
0 A
30
B
A 0
o A0
~
W ,,
Y s
20 10
0 0.001
0 0.0 1
0.1
1
Composite soil sample from a depth of 2-7 m. IONIC STRENGTH (M)
methanol solution containing 42% soil (w/w) was equilibrated for 1 h and then centrifuged at 1240g, the TCE was measured by direct GC column injection of an aliquot of the resultant supernatant, and then the procedure was repeated; and (b) a soil and water mixture (12% soil w/w) was equilibrated for 12 h and then centrifuged and the TCE measured by purge and trap of the resultant supernatant. The second methanol extraction following the first procedure did not yield any significant additional TCE, indicating that one methanol extraction was sufficient for the estimation of the total soil TCE. The efficiency of TCE recovery following the above described second procedure was 30%, based on the methanol extraction. Based on the above findings, the quantification of total TCE in contaminated soil samples was routinely determined as follows: a 1:3 ratio (w/w) mixture of methanol and soil was vortexed and placed in a sonicator bath at room temperature for 1 h. The mixture was shaken for 12 h and then centrifuged for 30 min at 1240g. TCE was quantified by direct GC column injection of an aliquot of the resultant supernatant and its TCE concentration was multiplied by the sum of the volume of methanol used and that of the water previously in the soil to determine total TCE in the soil sample. Data Analysis. Adsorption and desorption partition coefficients were determined based on the linear model, qe = K,C,, where K , is the partition coefficient (mL/g), C, is the solution-phase TCE concentration (ng/mL), and qe is the solid-phase TCE concentration (ng/g). The desorption partition coefficient was based on the partitioning obtained between the successively decreasing TCE concentration in the soil and the resulting equilibrium solution concentration in the daily, batch reequilibration experiment.
Results and Discussion Soil Properties. The properties of the soil used in this study are presented in Table I. Particle size analysis performed on a composite sample taken between depths of 2 and 7 m indicated that the soil was a silty sand (SM) according to the Unified Soil Classification System (ASTM-D2487) (10). However, the soil used for the adsorption and desorption experiments was primarily silty clay. The TCE concentrations in the soil samples used in this study varied between 1.6 and 9.9 pg/g (dry weight basis). Ionic Strength and pH Effects. No significant pH effect on the desorption of TCE was observed in a pH range from 3 to 13. The desorbed TCE varied from 29.1 % to 30.9% of the initial soil TCE (mean 30.0; SD = 1.2; n = 7 ) . This is expected since changes in pH should not affect the desorption of the nonpolar TCE. In general, pH changes do not affect sorption of neutral organic compounds such as chlorinated hydrocarbons (7). The increase of ionic strength to 0.1 M did not affect TCE desorption. However, decreasing TCE desorption 276
Environ. Sci. Technol., Vol. 25, No. 2, 1991
Figure 1. Effect of ionic strength on batch desorption of TCE due to NaCl (circles) and (NaPO,), (triangles).
s
t
0 0
,
,
,
5
10
15
, 20
, 25
, 30
I 35
TIME (hours) Figure 2. Kinetics of TCE batch desorption from a long-contaminated soil.
was observed when the ionic strength exceeded 0.1 M (Figure 1). While neutral molecules are generally less affected by salinity, their adsorption increases with increasing salt concentration (14,15). Conversely, desorption of these organic compounds should be negatively affected by increased salinity, as seen here for TCE. A possible cause for this behavior is flocculation of the clay particles resulting in reduced access to or from the micropores. Increasing NaCl concentration is known to initiate flocculation in clays (16). However, the same effect seen with the NaCl was observed when the system was dispersed with sodium hexametaphosphate (Figure 1). The dispersing agent reverses the clay edge charge and builds up the reverse charge to a level sufficient to prevent flocculation (17). Since the desorption was reduced in both the flocculated and the dispersed systems, micropore accessibility could not be responsible for limiting TCE desorption. The activity of TCE was probably reduced at and above an ionic strength of 0.1 M, as a result of the salting out effect. A t the lower ionic strength the desorption was unaffected because nonionic solutes have activity coefficients of approximately 1 for ionic strength of less than 0.1 M (18). Based on the fact that batch methods and relatively short times were used for the pH and ionic strength tests, the above described effects may be on the rapidly reversible TCE fraction and may have had very little effect on the slowly reversible fraction. Batch Desorption Kinetics. Daily readings of the percentage of TCE desorbed, was relatively constant at an average value of 25.3% (SD = 2.2; n = 7) over a 7-day equilibration period. When readings were taken at time intervals on the order of hours, the 25.3% steady-state desorption was approached at -22 h (Figure 2). This steady state is most likely an apparent equilibrium resulting from the inability to quantify minute increases in desorbed TCE. In systems where discrenment of small changes in sorption over extended periods is possible,
1
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,
,
,
,
,
,
,
1
2
3
4
5
6
7
0 0
1 0
8
100
200
300
TCE IN SOLUTION (ng/mi)
NO. OF EQUiLiBRATiONS Flgure 3. Removal of TCE by successive hourly (triangles) and daily (circles) batch reequilibrations
Figure 5. Estimation of the desorptive partiion coefficient of TCE. The slope of the regression line (K,) is equal to 19.7 mL/g.
25
2000
U
O Z 2 100 oo/
// 0
0
10
20
30
40
TCE IN SOLUTiON (ng/ml) Figure 4. Estimation of the adsorptive partiion coefficient of TCE. The slope of the regression line (K,) is equal to 11.7 mL/g.
desorption continues over periods of days to weeks (19). In batch studies where high water to soil ratios allow uniform mixing and adequate soil dispersion, shorter time periods are required for equilibration than under field conditions. Effect of Reequilibration. Simulations of successively decreasing soil TCE concentrations were achieved with the reequilibration experiments. A third of the TCE in the soil (1825 ng) was removed during the first daily equilibration. Subsequent TCE removal successively decreased for each additional day, with a total of 75% being removed after 7 days (3740 ng). The hourly equilibrations followed a similar pattern except that, overall, lower TCE removal was observed. After the first hour, 16% (800 ng) was removed and increased to 43% (2115 ng) after 7 h (Figure 3). Almost twice the TCE was removed in seven daily reequilibrations than in seven hourly reequilibrations. Apart from equilibration time, the concentration gradient established by successive replacements of the desorbing solution, also increased TCE desorption. After seven hourly reequilibrations, more TCE had desorbed than in 24 h during the first daily equilibration because of solution replacement. Although both the time and the frequent replacement of the desorbing solution enhanced TCE desorption, there is a greater temporal rather than concentration gradient dependence. Partition Coefficient. The isotherm for the adsorption of TCE by the silty clay soil used in this study is depicted in Figure 4. By use of linear regression techniques, the adsorption partition coefficient ( K ) was estimated as equal to 11.7 mL/g (SD = 4.4 mL/g; rH= 0.978; n = 5). Based on the successively decreasing soil TCE concentrations during the daily reequilibrations and the associated aqueous TCE concentrations, a desorptive partition coefficient of 19.7 mL/g (SD = 7.0 mL/g; r2 = 0.995; n = 7) was also obtained (Figure 5). The observed difference between these two partition coefficients derived from adsorption and desorption indicates that the system had
0
25
50
75
100
TIME (h) Flgure 6. Removal of TCE from a short continuous-flow soil column.
failed to reach equilibrium during the experiment. Attaining “true” equilibrium (which could take years) in the laboratory may be severely restricted by time constraints. In this study, the TCE concentration in the solution of the adsorption experiment leveled off after 3 days of mixing. It is conceivable, however, that TCE continued to adsorb, albeit at a very slow rate, resulting in undetectable difference of the aqueous TCE concentrations between consecutive TCE measurements. Estes et al. (20) observed an initial rapid TCE adsorption onto citrate/bicarbonate/dithionite-pretreated montmorillonite for a few days, followed by a slower adsorption rate, which continued without attaining equilibrium for mixing periods over 1 month. The sorption of nonpolar organic compounds in soils and sediments has been correlated to the organic carbon content of the sorbent (21,221. However, when the sorbent organic carbon content is below 0.1%, the sorbent organic carbon fraction is not a valid predictor of the partitioning of nonpolar organic compounds (22,231. In the later case, other sorbent properties such as specific surface area, type, and condition of the mineral surface may control the adsorption of nonpolar organic compounds. Recently reported data of TCE partitioning on organic sorbents, mineral oxides, and aquifer materials are compiled in Table 11. Based on these data, even for sorbents with relatively high organic carbon content, the organic carbon normalized partition coefficient (K0Jfor TCE varies by a factor of 3. The TCE partition coefficient (K,) for low organic carbon content sorbents is very low, ranging from 0.008 to 6.0 mL/g. The K , for TCE found in this study, when compared to the K , values for other sorbents with similar organic carbon content, is about 2- to 10-fold higher. A number of factors may contribute to the variability of Kp values for the low organic carbon content sorbents, such as (a) physical and chemical properties of the sorbent, (b) duration of adsorption and degree of mixing of sorbent/sorbate, (c) solids-to-liquid ratios, and (d) level of initial sorbate concentrations used. Environ. Sci. Technol., Vol. 25, No. 2, 1991 277
Table 11. Partition Coefficient ( K , ) of TCE for Organic Sorbents, Mineral Oxides, Clay, and Aquifer Materials" organic carbon, %
sorbent organic sorbents humic acid humic acid coated A1203 humic acid coated A1,0, mineral oxides and clay alumina (A1,OJ gibbsite [AI(OH),] goethite [FeO(OH)] montmorillonite aquifer materials surface soil soil (60% sand, 26.5% silt, 13.5% clay) soil (93% sand) soil (96% sand) soil (fine sand) soil (silty sand) soil (sandy clay) surface soil (Appalache Ap horizon) soil (5-m depth) soil (5-6-m depth; silty clay) NR, not reported. 560
I-
33.5 0.40 0.48
0.008
