Environ. Sci. Technol. 1992, 26, 1404-1410
Nonequilibrium Sorption and Aerobic Biodegradation of Dissolved Alkylbenzenes during Transport in Aquifer Material: Column Experiments and Evaluation of a Coupled-Process Model Joseph T. Angley,+ Mark L. Brusseau,*~tW. Lamar Miller,§ and Joseph J. Delfino§
Ecology and Environment, Inc., Fort Lauderdale, Florida 33304,Soil and Water Science Department, 429 Shantz Building, University of Arizona, Tucson, Arizona 8572 1, and Environmental Engineering Sciences Department, A. P. Black Hall, University of Florida, Gainesville, Florida 3261 1 We investigated the effect of simultaneous sorption and aerobic biodegradation on the transport of several dissolved alkylbenzenesin an aquifer material and evaluated the performance of a coupled-process transport model. First-order biodegradation rate constants decreased with decreasing number of C in the alkyl groups and exhibited a correlation with molecular structure as measured by molecular connectivity. For three series of isomers, the rate constant for the isomer with a substituent in the ortho position was smaller than the rate constants for the other isomers. Predictions obtained with the coupled-process model, wherein sorption was assumed to be rate limited, matched the breakthrough curves better than did predictions obtained with a model wherein sorption was assumed to be instantaneous. Accordingly, the assumptions upon which use of the model was based, e.g., biodegradation occurs only in solution and without a significant acclimation period, and can be simulated with a first-order equation, appear to be valid. Introduction Predicting the transport and fate of organic compounds in the subsurface requires an understanding of the operative physical, chemical, and biological processes and of the interactions among these processes. This is especially true for compounds such as the monoaromatic hydrocarbons, which can be affected by sorption and biodegradation (1-4). Coupled-process research involving the effects of sorption and degradation on the transport of organic compounds in the subsurface was initiated by soil scientists interested in the fate of pesticides (cf. refs 5-8). Such research has now been extended to compounds associated with industrial and commercial activities (cf. refs 9-11). A component of the coupled-process research has been the development of mathematical models designed to simulate the transport of solute that is affected by sorption and degradation. These models may be classified by the manner in which the processes of sorption, degradation, and transport are described (12). Many of the models first developed were based on the assumptions of instantaneous sorption and first-order degradation. While these models have been used successfully to simulate experimental data in a few cases (cf. ref 13), they have failed in other cases (cf. refs 7 and 8). Recently, coupled-process transport models have been presented that account for rate-limited sorption (14-17'). A version of these models that combines instantaneous and rate-limited sorption was used to simulate the transport and abiotic degradation of triazine pesticides (atrazine, cyanazine, simazine) in columns packed with soil (18,19). However, values of selected input parameters were adjusted to obtain optimal fits of the 'Ecology and Environment, Inc. t University of Arizona. University of Florida. 1404
Environ. Sci. Technol., Vol. 26, No. 7, 1992
model simulations to the data. These coupled-process transport models have rarely been evaluated by means of comparing independent predictions (i.e., where no input parameters are "fitted") to experimental data. The purpose of this study was to investigate the effect of simultaneous sorption and aerobic biodegradation on the transport of several dissolved alkylbenzenes in an aquifer material and to evaluate the performance of a coupled-process transport model that includes rate-limited sorption and first-order degradation. The miscible-displacement technique was used to generate breakthrough curves for the solutes. These data were then used for evaluating performance of the model. Materials and Methods Materials. The aquifer material used in this study was collected from a sandy, surficial aquifer that was contaminated by a spill of gasoline from a storage tank located at the Citrus Research and Education Center, Lake Alfred, FL. Samples of aquifer material were collected from slightly below the water table, approximately 1.2 m below the ground surface, at a point just inside the perimeter of the contaminant plume. The aquifer material was dried in an oven at 105 "C for 24 h. The aquifer material was then sieved; the less than 2-mm fraction was used for all experiments. Prior to being packed into the columns, the aquifer material was sterilized by being placed in an autoclave for 90 min for 3 consecutive days. Selected properties of the aquifer material are the following: clay, 1.8%; sand, 96.5%; silt, 1.7%; organic carbon, 0.015%; and pH (0.01 M CaClJ, 7.4. All solutions used in this study were composed of water collected from wells emplaced in the surficial aquifer. Contaminated water containing several alkylbenzeneswas collected from a well located within the contaminant plume; this water was used as influent for the miscibledisplacement experiments. Analysis of the water by GC/MS (EPA Method 602 modified by use of a 0.53mm-i.d., 30-m DB-1 column) revealed the following organic compounds: benzene (5), toluene (2.9), ethylbenzene (2.4), m-p-xylene (2.11, o-xylene (2.8), n-propylbenzene (0.8), isopropylbenzene (0.2), 2-ethyltoluene (1.5), 3-4-ethyltoluene (2.3), 1,2,3-trimethylbenzene (1.8), 1,2,4-trimethylbenzene (l.l), 1,3,5-trimethylbenzene (0.8); numbers in parentheses are the concentration in milligrams per liter and are the Co values for the column experiments. Note that the meta and para isomers of xylene were not resolved with the analytical procedures used in the study. Hence, the combination of these analytes is reported as m-pxylene. Likewise, 3-ethyltoluene and 4-ethyltoluene were not resolved and are reported as 3-4-ethyltoluene. The pH of the water ranged from 6 to 7, and the conductivity was approximately 300 pa. Total phosphate was 0.4 mg L-l, and nitrate was 0.29 mg L-l. The dissolved oxygen concentration (DO) was approximately8 mg L-l, which was sufficient to support aerobic biodegradation. Unconta-
00 13-936X/92/0926- 1404$03.00/0
0 1992 American Chemical Society
Table I. Conditions of Experiments expt bioactive L,cm 1 2 3
no yes yes
5 5 2.5
p,
g cm-3
0
u, cm h-'
1.82 1.82
0.29 0.29
12.2
a,cm
0.2
0.2
1.8
0.3
40.6 0.6
0.2
minated water was collected from a well located outside the plume; after being sterilized, this water was used to saturate the packed columns. Calcium chloride and tritiated water were used as nonsorbing tracers to characterize the hydrodynamic properties of the columns. Procedures. The apparatus and methods used for the miscible-displacement experiments were similar to those used by Lee et al. (20). The columns were Altex/Beckman preparative chromatography columns (No. 252-18) made of precision-bore borosilicate glass, with internal diameters of 2.5 cm. Adjustable shafts were used to provide packed-bed lengths of 2.5 and 5 cm. Bed supports on both ends of the columns consisted of a woven FEP Teflon diffusion mesh in contact with a porous Teflon filter disk, the pore diameters of which ranged from 30 to 60 pm. The diffusion mesh is designed to enhance radial distribution of influent solution and to reduce dispersion at the effluent end. The columns were packed in incremental steps with dry aquifer material to establish uniform bulk densities (see Table I). After completion of packing, uncontaminated well water was pumped through each column to establish complete saturation. The fractional volumetric water contents of the columns, measured by mass balance, are reported in Table I. To perform a miscible-displacement experiment, a column was connected to a single-piston HPLC pump (Gilson Medical Electronics Model 302) by means of stainless steel or Teflon tubing. The reservoirs for the influent solutions were Teflon gas-sampling bags (Alltech). The bags were evacuated to eliminate air and, thus, reduce volatilization. The bags were connected to the pump with stainless steel tubing. Breakthrough curves were obtained by pumping the influent solution through the column at a constant rate. Pumping was continued until steady-state effluent concentrations were achieved. For the experiments employing the 5-cm column, effluent fractions were collected manually in 1-mL crimp seal vials. These samples were either analyzed immediately or stored at 4 OC for later analysis; all samples were analyzed within 48 h. For the experiments employing the 2.5-cm column, effluent was analyzed immediately after being withdrawn through a low-volume in-line septa fitted to the column. Analyses of the hydrocarbons were performed by gas chromatography (Perkin Elmer Model 8410) with a flame ionization detector. Samples were concentrated by purge and trap with a Tekmar LSC/ALS system, using a modified version of EPA method 602. Separation was achieved with a 0.53-mm-i.d., 30 m long, fused-silica Megabore DB-1 (100% methylpolysiloxane) column (J & W Scientific). Chloride was analyzed with a chloridometer automatic titrator, and tritiated water was analyzed by radioassay (Searle Model Delta 300 LCS). Three sets of column experiments were performed in this study; conditions for all three are presented in Table I. The first experiment was designed to characterize the sorption and transport of the alkylbenzenes. Hence, bioactivity was prevented by placing the column after it was packed with aquifer material in an autoclave and by filtering all solution through a 0.2-pm filter. The second and third experiments were performed to evaluate the effect of biodegradation on transport, and therefore,
bioactivity was promoted. This was accomplished by inoculating the column after the autoclaving step prior to the experiment. Columns were inoculated by saturating the column with nonsterilized, contaminated groundwater, with contact times of at least 3 days. Sterilization by autoclaving, as well as by other techniques, affects soil properties and can alter sorptive behavior (21). Thus, the autoclaving-inoculation procedure was used instead of using untreated aquifer material to make conditions as similar as possible to the nonbioactive experiment. The biodegradation experiments were performed at two porewater velocities to investigate the effect of velocity on rate of biodegradation. All experiments were performed at approximately 23 "C. Preliminary batch microcosm experiments were performed to evaluate the potential for mass loss of the alkylbenzenes. A series of vessels containing aquifer material and unfiltered groundwater collected from the field site were incubated (20 f 1 "C) for 31 days. Samples were taken at 0,2,7,14,21, and 31 days and were analyzed for alkylbenzenes (GC),dissolved oxygen (YSI DO meter), and triphenylformazan (spectrophotometer). Production of triphenylformazan was used to measure dehydrogenase activity, which is used to assay the metabolic activities of microorganisms (22,23). Sterile controls were produced by addition of sodium azide (1.25 mg L-I). Mathematical Model. In this study, we will use a one-dimensional model that includes combined equilibrium and rate-limited sorption and first-order degradation. For these models, sorption dynamics is described with a bicontinuum or two-domain approach. In this approach, sorption is assumed to be essentially instantaneous for a fraction of the porous medium and rate limited for the remainder. Mass transfer of solute in the rate-limited domain is represented by a first-order equation based on the linear driving force approximation. Transport models incorporating this approach were presented by Selim et al. (24) and Cameron and Klute (25) and have been used with success by several researchers (cf. refs 26-31). The loss of solute mass by degradation is represented by a first-order equation. Although first-order kinetics has often been used on empirical grounds, it does have a theoretical basis (cf. refs 12 and 32). The first-order equation may describe the rate of biodegradation under conditions of steady-state biomass and low-substrate concentrations. In addition, it is assumed that the microbes are acclimated to the substrate (i.e., no acclimation period). The conditions under which our biodegradation experiments were performed approach those conditions required for first-order kinetics to be valid. The nondimensional governing equations for solute transport constrained by steady-state water flow, ratelimited linear sorption, and first-order degradation are (16)
PR
as* = -1 a2c* - ac* + (1- P)R ac* - [C* - qS* aT aT P ax2 a x (1 - P)R as*= o(C* - S*)- qS*
aT
(2)
where the following nondimensional parameters are defined as
c* = c/c,
s* = S,/[(l-
F)K,C,I
(34 (3b)
x =x/l
(3c)
T = tv/l
(34
Environ. Sci. Technol., Vol. 26, No. 7, 1992
1405
P = vl/D
(34
R = 1 + (p/B)Kp
(30
P = [1 + ( p / W ' K , I / R
(3g)
o
= kz(1 - P)Rl/V
E = PlZ/U + (PR - 1)ps11/u 7 = (1- P
)RP~Z~/~
YY.........
Toluene Data - Slmulatlon LEA model
(3h) (39 (38
where C is concentration of solute in the solution (M C3), S is the sorbed-phase concentration ( M M-l), Co is concentration of solute in the influent solution, t is time (T), x is distance (L), 1 is column length (L), p is bulk density of the porous medium (M ,T3), B is fractional volumetric water content of the packed column, u is average porewater velocity ( L T l ) , D is hydrodynamic dispersion coefficient (L2T1), Kp is equilibrium sorption constant (L3M I ) , F is fraction of sorbent for which sorption is instantaneous, kz is reverse sorption rate constant (Tl), and pi is the first-order degradation rate constant for the solution (l), equilibrium-sorbed (sl), and rate-limited-sorbed (s2) phases, respectively. The governing equations are solved with a finite-difference numerical approach under the following initial and flux-type boundary conditions: C*(X,O) = S*(X,O) = 0 (44
.g
0.8 -
, I Data
- Slmulatlon
,
0
1
2
3
4
5
(...)"
1,2,4 TMB Data Slmulatlon LEA model
-
The performance of the model that includes rate-limited sorption will be compared to that of a model wherein sorption is instantaneous, based on the local equilibrium assumption (LEA). This later model can be derived from the former by setting Szand S* equal to 0, and F and P equal to 1 in eqs 1-3. Data Analysis. Values for the Peclet number (i.e., dispersion coefficient, dispersivity) were obtained by fitting the LEA-based model to the breakthrough curves of the nonsorbing solutes; this was accomplished with a nonlinear, least-squares optimization program (33). Values for the retardation factors were determined by measuring the areas above the breakthrough curves obtained from experiment 1 (no biodegradation) (34-36). A nonlinear, least-squares optimization program using the nonequilibrium-based model (33) was used to determine values for P and w for the breakthrough curves obtained from experiment 1. Values for E, and thus pl (see eq 39, were determined in the following manner. Equations 1and 2 reduce to the following under conditions of steady state (5)
assuming that dispersive flux is minimal (Le., P > 25) and that biodegradation does not occur in the rate-limited sorbed phase, and where C,* is the steady-state relative effluent concentration. Integration of eq 5 leads to = - In C,*
(6) With the further assumption that only solution-phase solute degrades, we have V
pl = -(-
1
In C,*)
1406 Environ. Sci. Technoi., Voi. 26, No. 7, 1992
(7)
0.4
0.2 00
1
I
I
,
I
2
3
4
5
Pore Volumes Figure 1. Breakthrough curves from experiment 1 (no biodegradation) and simulations obtained with transport model that includes rate-limited sorption. 3-4 ET is 3-4-ethyltoluene; 1,2,4 TMB is 1,2,44rimethylbenzene.
Results and Discussion Preliminary Batch Biodegradation Experiment. For the nonsterile microcosms, there was a rapid decrease in the concentrations of all alkylbenzenes, with no evidence of an acclimation period. Concentrations for all solutes were below detection limits (0.5 pg L-l) by 31 days; sooner for many of the solutes. The decrease in alkylbenzene concentration was accompanied by a rapid decrease in DO (7.5-2.5 mg L-I) and a substantial increase in triphenylformazan (10-60 pg g-l). Conversely, there was little reduction in alkylbenzene concentrations (after accounting for sorptive uptake), no reduction in DO, and very little change in triphenylformazanfor the sterile controls. These results are consistent with mass loss via biodegradation. The mass loss of the alkylbenzenes was described well by a first-order rate equation. The rate constants for all compounds varied by less than a factor of 2, ranging from 0.2 to 0.3 day-l. Transport without Biodegradation. The breakthrough curves for chloride and tritiated water (not shown) were symmetrical and were invariant with velocity and column length. This behavior is representative of systems
Table 11. Parameter Values fast-velocity expt chemical benzene toluene o-xylene m-p-xylene ethylbenzene n-propylbenzene isopropylbenzene 2-ethyltoluene 3-4-ethyltoluene 1,2,3-trimethylbenzene 1,2,4-trimethylbenzene 1,3,5-trimethylbenzene
R 1.4 1.7 1.7 2.0 1.7 2.5 2.1
2.0 2.3 1.9 2.2 2.3
P
w
5
M , day-‘
NA” 0.68 0.72 0.62 0.74 0.60 0.71 0.70 0.64 0.70 0.66 0.66
NA 0.71 0.87 1.30 0.88 2.57 1.41 1.46 1.70 1.48 1.47 1.65
0.09 0.28 0.52 0.58 0.55 0.85 0.74 0.72 0.73 0.70 0.74 0.80
17.6 54.8 101.8 113.6 107.7 166.5 144.9 141.0 143.0 137.1 144.9 156.7
slow-velocity expt € PI, day-’ 1.05 6.0 2.2 12.5 2.3 13.1 3.5 19.9 2.75 15.6
NA, not applicable.
that are not influenced by transport nonideality (physical nonequilibrium) (20,37). The dispersivity values for the packed columns are reported in Table I. These values are similar to values obtained for columns packed with other sandy aquifer materials (20, 30, 38). The breakthrough curves for the alkylbenzenes were asymmetricaland exhibited delayed approaches to relative concentrations of 1 (see Figure 1 for examples). This behavior is consistent with behavior elicited by rate-limited sorption and concurs with the results of others (20, 30, 38-40), who have shown that sorption of organic compounds by aquifer materials can be rate limited. The optimized simulations obtained with the transport model that includes rate-limited sorption match the data well (see Figure 1). The parameter values determined from these analyses are reported in Table 11. The transport model based on the local equilibrium assumption for sorption could not simulate the data very well (see Figure 1). The magnitudes of the sorption rate constants are similar to those reported by others for similar solutes, sorbents, and conditions (20, 30, 31, 38, 42). Conditions where breakthrough curves for nonsorbing tracers are symmetrical but those for sorbing solutes are not are representative of systems influenced by sorptionrelated nonequilibrium (37). For many sorbents, an intrasorbent diffusion mechanism appears to be responsible for rate-limited sorption of organic compounds (31, 37, 40-44). This intrasorbent diffusion mechanism may involve domains located within the intraparticle pore space of microporous partjcles (40) or within the matrix of organic matter (31,37,43),or some combinationthereof (44). Solute located within the internal structure of the sorbent may be protected from biodegradation because of exclusion of bacteria (12,321. If so, it is possible that the assumption that only solution-phase solute is biodegraded would be valid. Transport with Biodegradation. Breakthrough curves for benzene, toluene, and o-xylene obtained from the biodegradation experiments are reported in Figure 2. These breakthrough curves, which are representative of all the results, exhibit the plateau in effluent solute concentration that is characteristic of transport under conditions of steady-state degradation. In all cases, the plateaus for the low-velocity experiments are lower than the plateaus of the high-velocity experiments. The decrease in mass degraded with increasing velocity is expected, considering that residence time decreases as velocity increases. The breakthrough curves obtained from the biodegradation experiments are compared in Figure 2 to the breakthrough curves obtained from the experiment
h b
0 0
e=
I
1
,
2
4
6
8
E 0
P
!?
c E al 0 C
s > -aal
b m
al
P 0.2
a
I
1
I
0.2 b b
0 0
2
0
NO D*ermdallon
H
D.gradmllon,
Y
* 40 cm/hr
,, 4
6
8
Pore Volumes Flgure 2. Selected breakthrough curves from all three experiments,
wherein biodegradation did not occur. Inspection reveals that the former are displaced to the right of the later. Thus, it is seen that degradation serves to delay the breakthrough of the solute. However, as will be shown below, this does not change the value of the retardation factor. Generally, degradation affects the magnitudes of the concentrations in the solution and solid phases but not Envlron. Sci. Technol., Vol. 26, No. 7, 1992
1407
200
150
0
Data
-
RegWDSiO"
& , = 9 5 8 X - 1709
*
".
r2=095 12
/ 0 1.5
2.5
2
I
3
3.5
4
X
Flgure 3. Relationship between first order biodegradationrate constant
(M,)and molecular connectivity ( X ) .
the equilibrium distribution between the two phases. The first-order biodegradation rate constants obtained from analysis of the fast-velocity breakthrough curves are reported in Table 11. The rate constants decrease with a decreasing number of alkyl-chain C atoms, with benzene having a rate constant lower than all of the alkylbenzenes. This order of rate constants for alkylbenzenes has been observed by others (45, 46). This apparent relationship between p1 and molecular structure can be evaluated quantitatively through the use of quantitative structureactivity relationships (QSAR). Several researchers have used QSAR to analyze biodegradation of organic compounds (cf. refs 47 and 48). Molecular connectivity (X), a measure of topological size and degree of branching, is a widely used molecular descriptor for QSAR analysis. The correlation between the plvalues obtained in this work and X is quite good and is shown in Figure 3. The apparent 1 ,
relationship between pl and molecular structure may result from the fact that the presence of an alkyl substituent presents microorganisms with two alternative pathways for biodegradation, oxidation of the ring or of the alkyl substituent, which can occur sequentially or simultaneously (4!9-51). Comparison of the rate constants for n-propylbenzene and isopropylbenzene reveals that the rate constant for the linear alkylbenzene is larger than that for the branched alkylbenzene. Similar data were reported by Bayona et al. (45),who showed that biodegradation of isomers increased when the phenyl group was closer to the end of the alkyl chain. Inspection of the data for the xylenes, the ethyltoluenes, and the trimethylbenzenes reveals that, in all three cases, the rate constant for the isomer with a substituent located in the ortho position is smaller than the rate constants for the other isomers. Thus, the placement of the substituents also appeared to influence the rate of biodegradation. Similar results have been reported by others (46, 52). The first-order biodegradation rate constants obtained from analysis of the slow-velocity breakthrough curves are reported in Table 11. Inspection shows that the patterns exhibited by the fast-velocity data are also exhibited by the slow-velocity data. Specifically, the rate constants decrease with decreasing number of C atoms in the functional groups, and o-xylene has a rate constant that is smaller than that of m-p-xylene. The rate constants obtained at the lower velocity are smaller, by a factor of 3-8, than those obtained at the higher velocity. The rate constants obtained from the batch biodegradation experiment are at least 1 order of magnitude smaller than those obtained from the slowvelocity column experiment. The apparent dependency
I
0.8 -
0.4
~
0 Data CI
l-
1 d)
2
3
Predictlon, LEA model
4
5
11
0
6
1
2
3
4
5
I 1,2,4 TMB 0 Data
0.6
0.4 -
0.4
'
1
0.2 I
0
Data
-
Prediction, RLS model
0
0.2 -
Prediction, LEA model I
/ 0
0
J
Toluene 0
1
I
1
I
I
2
3
4
5
Pore Volumes
0 6
L
I
0
1
2
3
4
5
6
Pore Volumes
Figure 4. Breakthrough curves from experiment 2 (with biodegradation) and predicted simulations obtained wlth two coupled-process transport modeis. LEA is model based on instantaneous sorption; RLS is model based on rate-limited sorption; 3-4 ET Is 3-4-ethyitoiuene; 1,2,4 TMB Is 1,2,4-trimethyibenzene. 1408
Environ. Sci. Technol., Voi. 26, No. 7, 1992
of pL1on residence time may result from the degree of mixing within and aeration of the reactor being dependent on flow rate. There would be little aeration and mixing in the batch reactor in comparison to the column system. In addition, the difference between rate constants for benzene and the alkyl-substituted compounds is greatest for the fast-velocity column experiment (e.g., factor of 8 between benzene and trimethylbenzenes) and least for the batch experiment (< factor of 2). This behavior may reflect the effect of residence time on the relative degradability or recalcitrance of benzene in comparison to the alkyl-substituted compounds. These data exemplify the “pseudo” or nonconstant nature of first-order biodegradation rate constants, which must be considered when first-order rate equations are used. Evaluation of Coupled-Process Transport Model. Examples of the predicted simulations obtained with two transport models are compared to the results of the fastvelocity biodegradation experiment in Figure 4. In all cases except that of benzene, the predictions obtained with the model that includes rate-limited sorption matched the data better than did the predictions obtained with the LEA-based model. Specifically, the nonequilibriumsorption-based model was able to simulate the asymmetry of the breakthrough curves. For benzene, however, the LEA-based model was sufficient. This is a result of the relatively small degree of sorption exhibited by benzene. When sorption is minimal (e.g., R < 1.5))nonequilibrium has minimal effect on transport and LEA-based models are, therefore, sufficient (30). The ability of the nonequilibrium-sorption-basedmodel to simulate the transport of the alkylbenzenes under conditions of biodegradationis especially encouraging since values for all parameters were obtained without curve fitting the specific data being predicted. To our knowledge, this is one of the first cases of validation of a coupledprocess transport model, incorporating rate-limited sorption and degradation, by means of independent predictions. It must be noted, however, that values for the biodegradation rate constants were obtained by massbalance analysis of the data being predicted and, thus, are not independent of the data being predicted. Another validation of a coupled-processtransport model, based on independent prediction, was recently reported (53);in this case, values for all parameters were obtained from independent experiments. On the basis of the successful performance of the model, it appears that the assumptions upon which the model is based are valid, at least for the systems used herein. One major assumption was that biodegradation could be modeled with a first-order equation. For our system, this appears to be true. Given the conditions of our experiments (i.e., low solute concentrations, no limiting substrates, acclimated bacteria), the validity of first-order kinetics was not surprising. Another assumption that is not inherent to the model but that was used to determine biodegradation rate constants was that biodegradation occurred only in the solution phase. The seeming validity of this assumption for cases such as ours is in accordance with recent experiments (cf. refs 54 and 55). Summary
Biodegradation under aerobic conditions was shown to have a significant effect on the transport of several dissolved alkylbenzenes in columns packed with an aquifer material. The degree of substitution and the placement of substituents influenced the values of the first-order biodegradation rate constant. The values of the rate constants were also dependent on the residence time as-
sociated with the experiment. A coupled-process transport model that includes rate-limited sorption and first-order degradation was used to predict successfully the breakthrough curves obtained from the column experiments. This indicates that the assumptions upon which use of the model is based, e.g., biodegradation occurs only in solution phase and without a significant acclimation period and can be modeled with a first-order equation, are valid at least for our particular system. This model should be useful for analyzing column experiments involving the transport of biodegradable compounds. Although these assumptions may also be valid for some field situations, extrapolation of these r e d t a to the field scale is, at present, problematic. Registry No. Benzene, 71-43-2; toluene, 108-88-3;o-xylene, 95-47-6; m-xylene, 108-38-3; ethylbenzene, 100-41-4; n-propylbenzene, 103-65-1; isopropylbenzene, 98-82-8; %ethyltoluene, 611-14-3; 3-ethyltoluene, 622-96-8; 4-ethyltoluene, 526-73-8; 1,2,3-trimethylbenzene,95-63-6; 1,2,4-trimethylbenzene,108-67-8; 1,3,5-trimethylbenzene, 620-14-4.
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Received for review October 14, 1991. Revised manuscript received February 12, 1992. Accepted February 13, 1992. This research was supported, in part, by the Citrus Research and Education Center, Institute of Food and Agricultural Sciences, University of Florida. The authors thank P. S. C. Rao, P . Nkedi-Kizza, and L. Lee for their assistance.
Laboratory Investigations on the Role of Sediment Surface and Groundwater Chemistry in Transport of Bacteria through a Contaminated Sandy Aquifer Martha A. Scholl US. Geological Survey, Water Resources Division, Mail Stop 439, 345 Middlefleld Road, Menlo Park, California 94025
Ronald W. Harvey' US. Geological Survey, Water Resources Division, US4, 325 Broadway, Colorado Colorado 80303-3328
The effects of pH and sediment surface characteristics on sorption of indigenous groundwater bacteria were determined using contaminated and uncontaminated aquifer material from Cape Cod, MA. Over the pH range of the aquifer (5-7), the extent of bacterial sorption onto sediment in uncontaminated groundwater was strongly pHdependent, but relatively pH-insensitive in contaminated groundwater from the site. Bacterial sorption was also affected by the presence of oxyhydroxide coatings (iron, aluminum, and manganese). Surface coating effects were most pronounced in uncontaminated groundwater (pH 6.4 at 10 "C). Desorption of attached bacteria (up to 14% of the total number of labeled cells added) occurred in both field and laboratory experiments upon adjustment of groundwater to pH 8. The dependence of bacterial sorption upon environmental conditions suggests that bacterial immobilization could change substantially over relatively short distances in contaminated, sandy aquifers and that effects caused by changes in groundwater geochemistry can be significant.
Introduction The transport of bacteria through porous media is a subject of current interest, having application in the fields 1410
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of bioremediation and public health. Strategies for in-situ remediation of contaminated aquifers require information on the transport of pollutant-adapted or genetically engineered microbes. Where drinking water supplies are derived from groundwater, transport of pathogenic bacteria from sewage and solid waste disposal sites is also of great concern. In sandy aquifers, the degree of bacterial transport can be governed largely by sorptive interactions with stationary grain surfaces. An important aspect of modeling microbial transport in porous media is how best to account for sorption processes (1-8). The DLVO (Derjaguin-Landau and Verwey-Overbeek) theory of colloid stability (9) describes the interaction of electrostatic and van der Waals forces in flocculation of colloidal suspensions and is often used, in part, to describe bacterial sorption in aqueous systems (10-13). Colloid filtration theory has been invoked in models describing experimental observations of bacterial transport within a sandy aquifer and in sand columns (2, 4 ) . Colloid filtration models quantify the physical mechanisms of particle contact with surfaces but use empirical coefficients to describe the effects of aqueous chemistry (14). Tobiason and O'Melia (15) modified the colloid filtration model to account for electrical double-layer (EDL) interaction forces; however,
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