Assessment of Bioavailability Using a Multicolumn System

Heating blocks (VWR Scientific Products Corp., Chicago, IL) were used to ..... Their conclusions were obtained from the initial CO2 production rate re...
1 downloads 0 Views 100KB Size
Download PDF
Environ. Sci. Technol. 2000, 34, 1506-1512

Assessment of Bioavailability Using a Multicolumn System XIANDA ZHAO AND THOMAS C. VOICE* Department of Civil and Environmental Engineering, Michigan State University, East Lansing, Michigan 48824

The bioavailability of sorbed naphthalene was studied with a surface soil using a multiple soil column system. This approach allowed monitoring contaminant concentrations in both the liquid and soil phases. Biodegradation and desorption parameters were determined in independent column experiments. A model in which desorption and biodegradation occur as sequential and independent processes successfully described the concentration of naphthalene in both phases. This result indicates that this organism only utilized liquid-phase contaminant, and direct soil-phase degradation did not occur. A previous study using the same organism concluded that direct soilphase degradation occurred on four different soils in completely mixed batch systems. The discrepancy between these reports appears to be caused by the different conditions existing in batch and column systems and the use of mineralization data to infer desorption rates in the earlier study. We conclude that if substrate depletion and mineralization rates are different, then desorption must be described using substrate concentration data from both the liquid and solid phases.

Introduction Bioavailability is a term used to describe the relative ease with which a substance can be used by organisms. In a system with only one contaminant phase (typically aqueous), utilization can normally be measured directly and is thought of as a characteristic of the organism and the solute. For systems containing multiple contaminant phases (e.g. aqueous and sorbed), utilization will be controlled not only by the organism but by contaminant mass-transfer processes as well (1, 2). In multiphase systems, the utilization rates should be expected to be different from those found in singlephase systems simply because the substance has multiple compartments in which it can reside and alternative pathways by which it can reach the organisms. Two sets of questions arise in this regard: (1) what are the compartments and pathways between compartments, and where does degradation occur, and (2) are there interactions between the processes or can each be treated independently? Biodegradation of organic contaminants from solids is often observed to depend on the substrate concentration in the aqueous phase, which is affected by sorption, desorption, and mass-transfer limitations (2). Sorption has been alternatively reported to increase, decrease, or have little effect on the bioavailabilities of individual organic compounds (1, 3). Weber et al. (4) reviewed the literature on the bioavailability of sorbed hydrophobic organic compounds including organic pesticides, herbicides, and halogenated hydrocarbons. They * Corresponding author phone: (517)353-9718; fax: (517)355-0250; e-mail: [email protected]. 1506

9

ENVIRONMENTAL SCIENCE & TECHNOLOGY / VOL. 34, NO. 8, 2000

concluded that even though organic chemicals react with organic and inorganic soil materials in many different ways, it is generally observed that bioavailability is inversely correlated with the amount of sorption by the soil. Ainsworth et al. (5) summarized several studies related to bioavailability of sorbed ionizable organic compounds such as aromatic acids and bases. They concluded that sorption can protect a compound from microbial attack and that this protection may be caused by inaccessibility of the micropores by microbes, surface stabilization against desorption of the compound, and reduction of aqueous-phase concentrations to levels below that necessary for microbial utilization. It has also been postulated that overall degradation rates may be determined by the desorption kinetics (6). Several studies have also reported that sorbed substrate can be utilized by the microorganisms at the rate faster than the desorption rate (7-16). The primary conclusion that can be drawn from this body of literature is that sorption-desorption and diffusion processes are very important in the biodegradation of organic compounds that can partition to solid materials, but that isolating these effects is experimentally difficult. An experimental approach that has been used to separate the effects of biodegradation and desorption is to compare systems with and without bioactivity. Harms and Zehnder (17) presented a study of the interactions between these processes using a model system with porous Teflon granules as a sorbent that simulates soil organic matter, 3-chlorodibenzofuran (3CDF) as a model contaminant, and Sphingomonas sp. strain HH19k as a test organism. The apparent half-saturation constant for substrate uptake was found to increase with increases in cell coverage in a packed column. They also found that the initial desorption rate increased with cell coverage until biodegradation was limited by oxygen transfer through the multiple layers of cells. They concluded that this phenomenon was the result of a localized reduction of substrate concentration at the surface of the cell producing an increased concentration gradient for desorption. Guerin and Boyd (11, 12, 18) studied the bioavailability of sorbed naphthalene to bacteria. Overall mineralization rates were measured in soil slurry systems with four soils, different soil:water ratios and two organisms: Pseudomonas putida ATCC 17484 and NP-Alk. They reported that ATCC 17484 effects a 7-fold enhancement of naphthalene desorption compared to NP-Alk. They also found that ATCC 17484 was positively chemotactic toward naphthalene, had a relatively hydrophobic cell surface, accumulated in high numbers in soil, and attachment was reversible. Strain NPAlk had a relatively hydrophilic cell surface and showed extensive and irreversible attachment to soil surfaces. They hypothesized that sorbed naphthalene was directly available to ATCC 17484 and complete desorption of bound naphthalene was facilitated by this organism, whereas strain NPAlk did not utilize the sorbed substrate or facilitate its desorption from soil. They also developed two empirical mathematical relationships to describe the mineralization rate of naphthalene in soil. Because of the complexity of the soil environment, mathematical modeling has been widely used to better understand biodegradation of contaminants during transport in porous media. To describe the biodegradation process per se, there are two major possibilities. The first is to assume that the contaminant can only be degraded in the aqueous phase, and the presence of the sorbent has no effect on degradation other than to establish the aqueous phase substrate concentration. This can be described mathematically using sequential and independent rate formulations of 10.1021/es991100b CCC: $19.00

 2000 American Chemical Society Published on Web 03/10/2000

the two processes. In the second scenario, the contaminant can be degraded directly in the solid phase as well as in the aqueous phase. This can be described mathematically using two degradation processes that are independent of the desorption process, that is, the rate coefficients are independent of the other processes occurring in the system. A further complication that must be considered is that degradation and desorption may not be independent, with the biomass serving to either enhance (e.g. due a surfactant effect) or suppress (e.g. due to blocking) desorption rates. To evaluate these inter-relationships it is necessary to design both an experimental approach and data analysis techniques capable of producing an independent evaluation of the two processes: biodegradation and desorption. We have approached this by (1) performing desorption-only and desorption/biodegradation column studies, (2) monitoring disappearance of the chemical from both the liquid and solid phases in order to separate the interactions between the two processes, and (3) employing a simple mathematical model which allows us to account for the effect of aqueous-phase degradation on desorption and to describe degradation in one or both phases.

Materials and Methods A surface soil, classified as Spinks loamy sand (SpcF), was collected at a depth of 1 ft from Ingham Co., MI. The soil was air-dried at 20 °C, sieved through a 0.425 mm sieve, and retained on a 0.250 mm sieve. The soil contained 78% sand, 17% silt, and 5% clay. The organic matter content was 1.9% and the pH was 4.2. During the experiment, phosphatebuffered saline (PBS) was passed through the soil, and the leachate pH was found to reach 6.92 after 30 pore volumes. The dry soil was placed in sealed containers and was sterilized by γ irradiation at a dosage of 2 Mrad in a cobalt-60 irradiator (Phoenix Memorial Laboratory, University of Michigan, Ann Arbor, MI). The soil was stored at 20 °C. Before the soil was used, microbial growth was examined by plating 0.1 g soil on a nutrient agar plate. There was no colony formation after 3 weeks incubation at 20 °C. Phosphate-buffered saline (PBS) solution was used as the mobile phase during the all column experiments. PBS solution contained 8.5 g of NaCl, 0.6 g of Na2HPO4, and 0.3 g of KH2PO4 per liter of D.I. water. The ionic strength was 0.16 M, and the pH was adjusted to 7.0 with addition of NaOH or HCl. Pseudomonas putida (ATCC 17484) was selected for the ability to grow on naphthalene as the sole source of carbon and energy. This aerobic, motile, gram-negative organism is also reported to be able to utilize sorbed naphthalene in a batch system (11). Five milliters of liquid ATCC 17484 culture with a density of 108 CFU/mL was used to inoculate 500 mL of high-buffer broth (2.0 g of NaCl, 3.0 g of (NH4)2HPO4, 1.2 g of KH2PO4, 3 mg of MgSO4, 1 mL of ferric quinate (19), 1 mL of vitamin solution (20), per liter of distilled water at pH 7.0) containing 200 mg/L of naphthalene. The culture was incubated at 20 °C with continuous mixing at 200 rpm. After the optical density (600 nm) of the solution increased to more than 0.2, which indicated that the culture had grown into the early stationary phase, cells were harvested by centrifugation at 1900×g for 20 min. The cells were rinsed with PBS three times to remove residual naphthalene and resuspended in PBS for use in the experiments. The naphthalene utilization rate under attached conditions was determined in a separate study (21). The maximum degradation rate (km) was 0.0211 L/h-ug Protein, and the half saturation coefficient (Ks) was 0.0496 mg/L. The cell attachment efficiency was 33% (21). Columns were constructed using stainless steel tubes (15 cm in length, 1 cm in I.D.) fitted with reducing unions containing 25 µm pore-size frits to prevent soil loss. A system

FIGURE 1. Schematic of multicolumn system. in which 12 identical columns could be run simultaneously was assembled, as shown in Figure 1. Columns were drypacked to an average value of bulk density of 1.32 g/mL, with a relative standard deviation of 2.45%, an average pore volume of 6.98 mL, with a relative standard deviation of 3.27%, and an average porosity of 0.47, with a relative standard deviation of 3.66%. Heating blocks (VWR Scientific Products Corp., Chicago, IL) were used to pasteurize (60 °C with retention time of 7 min) the column influent and effluent streams to reduce microorganisms from entering the columns and to reduce ATCC 17484 cells from entering the sampling loops. Plate counts confirmed a reduction of at least 2 orders of magnitudes in ATCC 17484 cells (from 105 CFU/mL to 103 CFU/mL). The length of tubing after the heating block was sufficient for the stream to cool to 20 °C before the columns or sampling loops. In the sorption mode of operation, a mechanical vacuum extractor pulled the water via 60-mL plastic syringes from the feed tank. During the desorption mode, the columns were fed by individual syringes using syringe pumps. To avoid sorptive losses of naphthalene, 1/16 in. OD stainless steel tubing (ID 1.27 mm) was used from the feed tank to the sample loops. The columns and fittings were sterilized by autoclaving at 120 °C for 30 min and dried at 105 °C for 48 h. The columns were packed with dry, sterilized soil and flushed with three pore volumes of carbon dioxide to replace the air in soil pores before water saturation to reduce air entrapment. Other components were sterilized with 70% ethanol for 1 h. Approximately 60 mL of sterilized PBS solution, which represented approximately eight pore volumes, was pulled thought the soil columns at a velocity of 8 cm/h during the saturation process. After saturation, the columns were removed from the system and capped. The total pore volume was determined by the weight difference before and after saturation. During the sorption mode, a velocity of 2.1 cm/h was used to feed a 1 mg/L naphthalene in PBS solution to the columns. The effluent concentration of naphthalene reached the influent level after 20 pore volumes, but the naphthalene VOL. 34, NO. 8, 2000 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

9

1507

feed was continued for 49 pore volumes to ensure uniform saturation. Four columns were removed from the system to determine the solid-phase naphthalene concentration. The mass of naphthalene retained in the soil columns was calculated based upon the integrated mass difference between the influent and the effluent of the columns. Two desorption experiments were conducted. In the first (desorption-only experiment), the columns were fed with naphthalene-free PBS solution at a velocity of 2.1 cm/h. In the second experiment (bioavailability experiment), 15 mL of ATCC 17484 culture (2.4 mg protein/L) was inoculated into each column at a velocity of 10 cm/min after the sorption mode; the columns were then fed with naphthalene-free PBS solution at a velocity of 2.1 cm/h. Two types of samples were collected. Effluent was obtained from the sampling loops, while influent was collected from the feed tank. Soil samples were obtained by sequentially removing and sacrificing columns from the system. This was accomplished by removing the end fittings, pushing the soil out and dividing it into six approximately equal segments that were transferred to preweighed vials containing 15 mL of methanol. The vials were sealed with a Teflon-faced cap and turned end-over-end at 6 rpm for 48 h. The soil was allowed to settle for 24 h, and 1 mL of supernatant was removed for analysis. The concentration of naphthalene was determined by high-pressure liquid chromatography (HPLC, Perkin-Elmer Series 200 LC) using a reverse-phase C-18 column and a mobile phase consisting of 80% acetonitrile and 20% water at a flow rate of 1 mL/min. Naphthalene was detected by UV absorption at 220 nm. The detection limit was 1.2 µg/L. To describe the kinetics of desorption, we used a two-site sorption-related nonequilibrium (equilibrium/kinetic) transport model (22, 23). In this formulation, the solid phase is divided into two types of sites: type I sites where sorption is at equilibrium and type II sites where sorption is kinetically controlled. For steady-state flow in a uniform system, the transport-degradation equation for the liquid phase can be written as

(

1+

)

FfKd ∂c ∂2c ∂c )D 2-υ θ ∂t ∂x ∂x FKd kmXc FfKd [(1 - f)Kdc - s2] µ c (1) θ Ks + c θ s

and the kinetic sorption/desorption process can be described as

F

∂s ) Fk[(1 - f)Kdc - s2] - Fµss2 ∂t

(2)

where s1 and s2 are solid-phase concentration for type I and type II sites, respectively; c is the water phase concentration of the solute, f is the fraction of exchange sites assumed to be at equilibrium, Kd is the sorption distribution coefficient, F is the soil bulk density, k is a first-order kinetic desorption rate coefficient, D is the dispersion coefficient, v is the average pore-water velocity, θ is the volumetric water content, km is the maximum degradation rate from the liquid phase, Ks is the half saturation coefficient, X is the biomass concentration, and µs is a first-order biodegradation coefficient in the solid phase (24). For the desorption only case, the biodegradation parameters in eqs 1 and 2 will be set to zero. Analytical solution can be accomplished using the technique of van Genuchten and Wagenet (22). The nonlinear least squares inversion program, CXTFIT, based upon this model was developed by U.S. Salinity Laboratory (25) and was used for parameter estimation and prediction in this study. To verify the numerical solution, a second numerical model was developed 1508

9

ENVIRONMENTAL SCIENCE & TECHNOLOGY / VOL. 34, NO. 8, 2000

using a modeling software package (ModelMaker, Cherwell Scientific Publishing, Oxford, Great Britain). The results from both programs were examined for several system conditions. No discrepancies could be found in either liquid- and solidphase concentrations. Parameter Estimation. The average pore-water velocity (v) was determined from the volumetric flow rate (Q), water content (θ), and the cross-sectional area of the column (A)

v)

Q Aθ

(3)

To determine the dispersion coefficient (D), three identical soil columns were prepared and saturated using the same procedure as described in previous section. Tritiated water (3H2O with activity of 1 × 105 DPM/mL) was fed to the columns at 2.1 cm/h via the syringe pumps. After the full breakthough of 3H2O was observed, clean PBS was fed to the columns, and the effluent samples were collected every 0.1 pore volumes. The concentration of the 3H2O was measured by liquid scintillation counting. The dispersion coefficient was estimated from the effluent breakthrough curves using a leastsquares optimization procedure in CXTFIT. It was found to be 0.4 cm2/h which was one magnitude higher than the molecular diffusion coefficient of 0.036 cm2/h for naphthalene (26). Therefore, adjustment for the different molecular diffusivities of the 3H2O and naphthalene was not necessary. The value of Kd was calculated from following equation

Kd )

Me M s Ce

(4)

where Me is the average naphthalene mass calculated from the effluent curves of desorption-only columns during the desorption (µg); Ms is average soil weight in the columns (kg); and Ce is average CT concentration in the effluent at the end of the sorption phase (µg/L). The nonlinear least squares inversion program, CXTFIT, was utilized to determine the value of k and f from the desorption break through curves of nonbioactive columns. The value of km and Ks was determined from data collected from the nonsorbed bioactive soil columns (21). The prediction of the solid- and liquidphase naphthalene concentration in the bioactive soil column was accomplished by using a modeling software package (ModelMaker, Cherwell Scientific Publishing, Oxford, Great Britain).

Results and Discussion Sorption and Measurement Solid-Phase Naphthalene. After completion of the sorption phase for the desorption-only experiments, four columns were taken offline for measurement of solid-phase naphthalene concentrations. The average total mass of naphthalene retained in the soil columns was 127.5 µg, with a relative standard deviation of 6.76%. As can be seen in Figure 2, the distribution of naphthalene along the length of the columns is relatively uniform. The extraction recovery efficiencies for all four columns were determined based upon the naphthalene mass as calculated from the influent and effluent data, and the measured value from the extraction (Table 1). Average recovery was 93.5%, so all extraction results presented in this study were not adjusted. During the desorption phase, time-dependent naphthalene concentrations were monitored in both the soil and effluent by periodically sacrificing columns and collecting liquid samples from the remaining columns. The naphthalene mass released during the desorption period was calculated using these data. An average value of 128.1 µg was found, which is indistinguishable from the value determined during the sorption period (127.5 µg). The sorption distribution coefficient (Kd) was estimated as 4.72 L/kg using eq 4. The

FIGURE 3. Measured and fitted naphthalene concentrations in the effluent of desorption-only soil columns during the desorption period. FIGURE 2. Solid-phase naphthalene distributions at the end of the sorption period during the desorption-only experiments.

TABLE 2. Desorption Model Parameters

TABLE 1. Naphthalene Extraction Recovery Efficiency

symbol

value

v D Fb θ Kd R

2.1 cm/h 0.40 cm2/h 1.31 kg/L 0.48 4.72 L/kg 13.85

f k

0.82 0.008 1/h

Independent Parameter

column no.

naphthalene mass calculated from the influent and effluent (µg)

naphthalene mass measured from extraction (µg)

extraction recovery efficiency (%)

1 3 6 12

126.6 123.2 126.6 122.1

113.1 114.8 115.3 113.2

89.4 97.0 91.1 96.4

sorption distribution coefficient was also determined using a batch method (27) and found to be linear but somewhat higher (6.49 L/kg, r 2 ) 0.955). Differences between batch and column techniques have been discussed by several researches (28-31). Several possible reasons, including mixing differences in the two systems, the effects of immobile water, loss of sorbent particle from the column, slow sorption process, and differences in the soil/water ratio between two systems, have been proposed. However, there is no clear consensus in the literature on this issue. We used the value determined from the column data in all further analyses. After the hydrological parameters including average porewater velocity (v) and hydrodynamic dispersion coefficient (D) were determined, and the equilibrium desorption coefficient (Kd) was calculated from eq 4, the mass-transfer desorption rate constant (k) and fraction of instantaneous sorption sites (f) were determined from the elution curve of desorption-only columns using the two-site model (Figure 3). The input parameters and the results of the estimation procedures are shown in Table 2. It can be noted that only 18% of the sites exhibited rate-limited desorption. Using the parameters estimated from the liquid-phase data, solid-phase concentrations were predicted, and good agreement was found for most of the desoption period (Figure 4). Some underestimation of naphthalene concentrations was observed late in the desorption process, although the trend in the profile was still similar. This may be a limitation in the two-site model, which assumes a constant and uniform masstransfer coefficient, whereas the process is perhaps better described as a variety of rates relating to the complexity of the soil environment (32). A distributed mass-transfer compartment model has been developed to describe the desorption process for highly aged soils (33). Given the high degree of correlation between the prediction and the data

average pore-water velocity hydrodynamic dispersion coefficient bulk density porosity equilibrium sorption/desorption coefficient retardation factor

Estimated from Curve Fitting fraction of instantaneous sorption sites mass-transfer desorption rate constant

TABLE 3. Comparison of Measured and Predicted Naphthalene Mass in the Desorption/Biodegradation Soil Columns comparison to measured value (%) simulation 1/10 simulation 1/20 no. of pore measured biodegradation biodegradation volume (µg) simulation rate rate 3.2 6.6 10.1 13.7 16.4 20.1

10.9 8.7 6.1 5.7 5.8 4.7

107.7 109.8 130.8 112.4 93.2 94.1

114.6 118.1 142.5 123.2 103.9 105.5

116.2 119.8 144.6 125.0 105.4 107.1

for most of the period, we did not find justification for the use of a more complex model to address our objectives. After the desorption rate parameters were determined from the desorption-only columns, the soil columns were repacked and fed with naphthalene contaminated PBS for 50 pore volumes. Four columns were again extracted, and it was verified that uniform solid-phase naphthalene concentrations were reached. Fifteen milliters of stationary cells (ATCC 17484, 2.4 mg protein/L) were introduced into each column at 10 cm/min. The desorption process was initiated by feeding naphthalene-free PBS (air saturated) at 2.1 cm/h. The first set of effluent samples was collected when 3.2 pore volumes of PBS had passed through after inoculation. No naphthalene was found in four of the columns, while the remaining four columns had from 0.031 to 0.27 mg/L, corresponding to 4% to 36% of expected initial liquid-phase naphthalene concentration (at the end of the sorption phase) in the columns (Figure 5). This suggests significant degradation capability had been established in all columns, as the VOL. 34, NO. 8, 2000 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

9

1509

FIGURE 4. Measured and predicted solid-phase naphthalene concentrations in the desorption-only columns during desorption period. The solid line represented the prediction from the modeling results except the line for the zero pore volume which was calculated from the average values of extraction results.

FIGURE 5. Measured and predicted liquid-phase concentrations of desorption and biodegradation columns during the desorption period. effluents of the desorption-only columns had not dropped appreciably at the same point (Figure 3). The variation in bioactivity is likely the result of differences in inoculation efficiency, acclimation, or the distribution of organisms relative to the flow paths. Liquid- and solid-phase naphthalene concentrations were predicted using a numerical solution of the model presented above with the assumption of that only liquid-phase naphthalene can be degraded. The dispersion coefficient was 1510

9

ENVIRONMENTAL SCIENCE & TECHNOLOGY / VOL. 34, NO. 8, 2000

FIGURE 6. Measured and predicted naphthalene mass retained in the soil columns during the desorption period in the desorption and biodegradation experiments. obtained from the tritiated water tracer study. The desorption rate parameters were estimated from the desorption-only columns. The biodegradation parameters were estimated from a nonsorptive sand column (21). Due to the very fast degradation rate of naphthalene, the effluent naphthalene concentration was predicted to reach zero after 0.5 pore volumes (Figure 5), consistent with our observations in four of the columns. To evaluate whether the higher levels in the other four columns could result from slower degradation kinetics, we also simulated profiles using utilization rates 1/10 and 1/20 of the estimated value (Figure 5). These curves provided a reasonable representation of the data. Predictions of the solid-phase data are shown in Figure 6. Good agreement was found. It can also be seen that differences in the biodegradation rate are important only during the initial period, where the substrate originates in either the liquid phase or equilibrium sorption sites. The system clearly becomes desorption limited after approximately 2 pore volumes and thereafter is unaffected by the degradation rate. The soil-phase naphthalene concentration profiles along the length of the columns, both initially and at 3.2 pore-volumes, are presented in Figure 7 along with the model predictions using the measured liquid-phase degradation rate. Again, very good agreement was found. Adding a solid-phase degradation term to the model resulted in significantly lower solid-phase naphthalene concentrations than were observed experimentally. We draw two conclusions from these results. First, the data in this system were well-described by a simple model involving sequential desorption and degradation processes and only liquid-phase biodegradation. Second, the very different length profiles of solid phase naphthalene in the two systems suggest that contaminant removal from the soil in a desorption-only system is dependent on the number of pore volumes flushing the system, whereas the biodegradation system is time dependent. In the later case, performance is not dependent on flushing once activity is established. The only flow necessary is that required to sustain activity,

FIGURE 7. Measured and predicted solid-phase naphthalene concentrations in the desorption/biodegradation columns. such as delivery of nutrients or electron acceptors/donors. Similar results have been found by other investigators. Gamerdinger et al. (34) summarizes the results of several pesticide sorption and degradation soil column studies. They did not find the clear dependencies of biodegradation on sorption and conclude that liquid-phase degradation predominates. Gamerdinger et al. (34) also pointed out the importance of using the extraction process to determine the remaining target chemical at end of the soil column experiment for the biodegradation. Ogram et al. (35) also reported that the liquid-phase 2,4-dichlorophenoxyacetic acid (2,4D) was utilized by both suspended and attached microorganisms in a soil slurry system while the sorbed 2,4-D was not directly utilized. Estrella et al. (36) reported that the overall biotransformation rate could be described using a liquidphase degradation coefficient although the rate parameters calculated from a batch experiment were significantly different from a values determined from the column experiment. Recently, we studied biotransformation of carbon tetrachloride (CT) by Pseudomonas sp. Strain KC in soil columns. Desorption and biotransformation could be described as sequential processes (24). Guerin and Boyd (11) reported that ATCC 17484 was able to utilize sorbed naphthalene directly from soil. Their conclusions were obtained from the initial CO2 production rate results in a batch system. Our organism/substrate combination was selected based on this report to evaluate how this suggested result would affect performance under transport conditions. However, we did not find evidence for solid-phase degradation. These apparently conflicting results may derive from experimental differences between the two studies. Guerin and Boyd conclude that direct mineralization of sorbed naphthalene occurred in all four tested soils for ATCC 17484. Using an empirical mathematical relationship, a curve-fitting technique was used to estimate the parameters in the equations. This formulation, however, does not clearly delineate desorption and biodegradation processes, and the results should be considered as an overall description of the system. Furthermore, the result is based on CO2 production data, with the implicit assumption that this directly reflects the naphthalene degradation rate. In the current study, the disappearance of naphthalene was measured directly. In a separate study, the mineralization of the naphthalene by ATCC 17484 was found to be a sequential process, in which the naphthalene was rapidly converted to less hydrophobic intermediate(s) and then the CO2 production occurred at a much slower rate, following a zero-order reaction (21). Since

naphthalene desorption is controlled by the solid:liquid naphthalene concentration gradient, we feel that our approach more accurately reflects naphthalene dynamics in the system. The different conclusions in this study may also result from the use of columns rather than batch systems. When initial mineralization rates or degradation rates are studied in well mixed-batch systems, desorption will be dominated by release from equilibrium sorption sites. Desorption in columns involves both equilibrium and rate-limited processes, which are described in our model formulation. Due to the very fast degradation rate in this study, the equilibrium sites were rapidly exhausted after first three pore-volumes (approximately 20 h), and the data primarily reflects ratelimited desorption. Nonequilibrium desorption is commonly found in contaminated soils and frequently dominates transport and remediation processes (37). Under such conditions, the contaminant may be trapped within soil organic matter or micropores and direct access by microorganisms presumably impossible. The natural aging process will further increase these restrictions (38-42). For the very fast or equilibrium desorption sites, the contaminant will likely reside on the outer layer of the soil particle, where it may be accessible to the organisms.

Acknowledgments This work was supported in part by the U.S. Environmental Protection Agency under the Great Lakes and Mid-Atlantic Center for Hazardous Substance Research.

Nomenclature A

cross-sectional area of the column

c

water phase concentration of the solute

Ce

average CT concentration in the effluent at the end of the sorption phase

D

dispersion coefficient

f

fraction of exchange sites assumed to be at equilibrium

k

first-order kinetic desorption rate coefficient

Kd

sorption distribution coefficient

km

maximum degradation rate from the liquid phase

Ks

half saturation coefficient

Me

average naphthalene mass calculated from the effluent curves of desorption-only columns during the desorption

Ms

average soil weight in the columns

Q

volumetric flow rate

s1

solid-phase concentration for type I

s2

solid-phase concentration for type II sites

X

biomass concentration

µs

first-order biodegradation coefficient in the solid phase

v

average pore-water velocity

θ

volumetric water content

F

soil bulk density

Literature Cited (1) Rao, P. S. C.; Bellin, C. A.; Brusseau, M. L. In Sorption and Degradation of Pesticides and Organic Chemical in Soil; Linn, VOL. 34, NO. 8, 2000 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

9

1511

(2)

(3) (4)

(5)

(6) (7) (8) (9) (10) (11) (12)

(13) (14) (15) (16) (17) (18) (19) (20) (21)

D. M., Carski, T. H., Brusseau, M. L., Chang, F. H., Eds.; Soil Science Society of America and American Society of Agronomy: Madison, WI, 1993; pp 1-26. Alexander, M.; Scow, K. M. In Reactions and Movement of Organic Chemicals in Soils; Sawhney, B. L., Brown, K., Eds.; 1989 Soil Science Society of America and American Society of Agronomy: Madison, WI, 1989; pp 243-269. van Loosdrecht, M. C. M.; Lyklema, J.; Norde, W.; Zehnder, A. J. B. Microbiol. Rev. 1990, 54, 75-87. Weber, J. B.; Best, J. A.; Gonese, J. U. In Sorption and Degradation of Pesticides and Organic Chemicals in Soil; Soil Science Society of America and American Society of Agronomy: 1993; pp 153196. Ainsworth, C. C.; Frederickson, J. K.; Smith, S. C. In Sorption and Degradation of Pesticides and Organic Chemical in Soil; Linn, D. M., Carski, T. H., Brusseau, M. L., Chang, F. H., Eds.; Soil Science Society of America and American Society of Agronomy: Madison, WI, 1993; pp 125-144. Scow, K. M. In Sorption and Degradation of Pesticides and Organic Chemicals in Soil; Soil Science Society of America and American Society of Agronomy: 1993; pp 73-114. Marshman, N. A.; Marshall, K. C. Soil Biol. Biochem. 1981, 13, 127-134. Ortega-Calvo, J.-J.; Saiz-Jimenez, C. Appl. Environ. Microbiol. 1998, 64, 3123-3126. Griffith, P. C.; Fletcher, M. Appl. Environ. Microbiol. 1991, 57, 2186-2191. Laor, Y.; Strom, P. F.; Farmer, W. J. Water Res. 1999, 33, 17191729. Guerin, W. F.; Boyd, S. A. Appl. Environ. Microbiol. 1992, 58, 1142-1152. Guerin, W. F.; Boyd, S. A. In Sorption and Degradation of Pesticides and Organic Chemical in Soil; Linn, D. M., Carski, T. H., Brusseau, M. L., Chang, F. H., Eds.; Soil Science Society of America and American Society of Agronomy: Madison, WI, 1993; pp 197-208. Calvillo, Y. M.; Alexander, M. Appl. Microbiol. Biotechnol. 1996, 45, 383-390. Tang, W. C.; White, J. C.; Alexander, M. Appl. Microbiol. Biotechnol. 1998, 49, 117-121. Crocker, F. H.; Guerin, W. F.; Boyd, S. A. Environ. Sci. Technol. 1995, 29, 2953-2958. Guerin, W. F.; Boyd, S. A. Water Res. 1997, 32, 1504-1512. Harms, H.; Zehnder, A. J. B. Appl. Environ. Microbiol. 1995, 61, 27-33. Guerin, W. F.; Boyd, S. A. In A Tri-Service Workshop on: Bioavailability of Organic Contaminants in Soil and Sediments; Monterey, CA, 1995. Blakemore, R. P.; Maratea, D.; Wolfe, R. S. J. Bacteriol. 1979, 140, 720-729. Wolin, E. A.; Wolin, M. J.; Wolfe, R. S. J. Biol. Chem. 1963, 238, 2882-2886. Park, J.-H.; Zhao, X.; Voice, T. C. Environ. Prog., submitted for publication.

1512

9

ENVIRONMENTAL SCIENCE & TECHNOLOGY / VOL. 34, NO. 8, 2000

(22) van Genuchten, M. T.; Wagenet, R. J. Soil Sci. Soc. Am. J. 1989, 53, 1303-1310. (23) Brusseau, M. L.; Jessup, R. E.; Rao, P. S. C. Water Resour. Res. 1992, 28, 175-182. (24) Zhao, X.; Szafranski, M. J.; Maraqa, M. A.; Voice, T. C. Environ. Toxicol. Chem. 1999, 18, 1755-62. (25) Toride, N.; Leij, F. J.; van Genuchten, M. T. U.S. Salinity Laboratory: Riverside, CA, 1995. (26) Schwarzenbach, R. P.; Gschwend, P. M.; Imboden, D. M. John Wiley & Sons: New York, 1993; p 2. (27) Voice, T. C.; Pak, D.; Zhao, X.; Shi, J.; Hickey, R. F. Water Res. 1992, 26, 1389-401. (28) MacIntyre, W. G.; Stauffer, T. B. Chemosphere 1988, 17, 21612173. (29) Lion, L. W.; Stauffer, T. B.; MacIntyre, W. G. J. Contam. Hydrol. 1990, 5, 215-234. (30) Piatt, J. J.; Backhus, D. A.; Capel, P. D.; Eisenreich, S. J. Environ. Sci. Technol. 1996, 30, 751-760. (31) Maraqa, M. A.; Zhao, X.; Wallace, R. B.; Voice, T. C. Soil Sci. Soc. Am. J. 1998, 62, 142-152. (32) Weber, W. J. J.; Mcginley, P. M.; Katz, L. E. Environ. Sci. Technol. 1992, 26, 1955-1962. (33) Culver, T.; Hallisey, S.; Sahoo, D.; Deitsch, J.; Smith, J. Environ. Sci. Technol. 1997, 31, 1581-1588. (34) Gamerdinger, A. P.; Dowling, K. C.; Lemley, A. T. In Sorption and Degradation of Pesticides and Organic Chemical in Soil; Linn, D. M., Carski, T. H., Brusseau, M. L., Chang, F. H., Eds.; Soil Science Society of America and American Society of Agronomy: Madison, WI, 1993; pp 115-123. (35) Ogram, A. V.; Jessup, R. E.; Ou, L. T.; Rao, P. S. C. Appl. Environ. Microbiol. 1985, 49, 582-587. (36) Estrella, M. R.; Brusseau, M. L.; Maier, R. S.; Pepper, I. L.; Wierenga, P. J.; Miller, R. M. Appl. Environ. Microbiol. 1993, 59, 4266-4273. (37) Brusseau, M. L.; Rao, P. S. C. Crit. Rev. Environ. Controls 1989, 19, 33-99. (38) Steinberg, S. M.; Pignatello, J. J.; Sawhney, B. L. Environ. Sci. Technol. 1987, 21, 1201-1208. (39) Hatzinger, P. B.; Alexander, M. Environ. Sci. Technol. 1995, 29, 537-545. (40) Allard, A.-S.; Hynning, P.-A.; Remberger, M.; Neilson, A. H. Appl. Environ. Microbiol. 1994, 60, 777-784. (41) Tang, J. X.; Carroquino, M. J.; Robertson, B. K.; Alexander, M. Environ. Sci. Technol. 1998, 32, 3586-3590. (42) Robinson, K. G.; Farmer, W. S.; Novak, J. T. Water Res. 1990, 24, 345-350.

Received for review September 27, 1999. Revised manuscript received January 18, 2000. Accepted January 19, 2000. ES991100B