Inhibition of Phenanthrene Mineralization by Nonionic Surfactants in Soil-Water Systems Shonall Laha and Richard G. Luthy"
Department of Civil Engineering, Carnegie Mellon University, Pittsburgh, Pennsylvania 15213 The biodegradation of [14C]phenanthreneby a polycyclic aromatic hydrocarbon (PAH) degrading inoculum was studied in water and in soil-water systems with nonionic surfactants. The purpose of surfactant addition was to assess the effect on biodegradation of liquid-phase PAH solubility enhancement via surfactant micellization. The nonionic surfactants were selected on the basis of their demonstrated PAH-solubilizing capacities in soil-water systems, viz., an alkylethoxylate, C12E4,and alkylphenol ethoxylate surfactants, CBPE9.5and C9PE10.5. In the presence of surfactants at concentrations that resulted in aqueous-phase cmc or micelle formation, the mineralization of [14C]phenanthrene was substantially inhibited. This inhibition was reversible on diluting the surfactant to a concentration below that resulting in micellization. Sub-cmc levels of surfactant in soil-water systems did not appear to have an inhibitory effect on phenanthrene mineralization, but neither did such doses serve to enhance the rate of degradation. Companion studies with [14C]glucose, phenanthrene, and surfactants suggest that the supra-cmc inhibitory effect on biodegradation is not a toxicity phenomenon, per se, of surfactant or micellized PAH, or a consequence of the nonionic surfactant being used as preferential substrate. Chemical modeling of the phase partitioning of phenanthrene between water, soil, and micelle shows that micellization results in a decrease in the equilibrium free aqueous phenanthrene concentration insufficient to explain the microbial inhibition observed. Additional tests are required to assess the role of surfactants as competitive substrates, the interaction of nonionic surfactant with membrane protein, and the bioavailability of micellized PAH and to determine whether the inhibitory effect is specific to the nonionic surfactants employed in this work. W
Introduction Polycyclic aromatic hydrocarbon (PAH) compounds are nonpolar hydrophobic molecules comprising two or more fused benzene rings, with aqueous solubilities ranging from -32 mg/L for naphthalene to 98%, Aldrich). The PAH compounds used for the initial solubilization studies were phenanthrene, anthracene, and pyrene (Table I). Mineralization experiments described in this paper were performed with phenanthrene. Both the solubilization and mineralization experiments used radiotracer techniques. 14C-Labeled PAH compounds and [14C]glucosewere obtained from Amersham Corp., Arlington Heights, IL. PAH-spiking solutions, consisting of mixtures of labeled and unlabeled PAHs, were prepared in methanol and used within 2 h. The spiking solutions for glucose were prepared in deionized water. Analysis of [14C]compoundwas taken as representative of the behavior of the compound as a whole. Samples were counted for 14Con a Beckman LS 5000 TD liquid scintillation counter, using the H# quench monitoring technique with automatic quench compensation, and computer data logging. The soil used was a Hagerstown silt loam collected from the A horizon of the Agricultural Experimental Station, Pennsylvania State University. The soil was air-dried and screened to pass a US. standard no. 10-mesh (2 mm) sieve. The fraction organic carbon in the soil was determined to be 1.5% by the Walkley-Black method (31). The PAH doses for the various tests with soil were designed to attain an initial aqueous-phase concentration equal to the PAH solubility limit in the absence of surfactant. Thus,the total amount of PAHs in the soil-water systems was the sum of the aqueous-phase PAHs and the PAHs sorbed on the soil in equilibrium with PAH near aqueous solubility (32). The surfactants used in this work were three nonionic polyethoxylates described in Table 11. They are a subset of those selected initially on the basis of a literature survey of surfactant-aided soil-washing and biodegradation studies; solubilization tests showed that these nonionic surfactants exhibited best PAH-solubilizing capacities in soil-water systems at doses around 1% (v/v) (32). Laboratory-grade surfactants Brij 30 and Triton X-100 were obtained from Aldrich Chemical Co., and Tergitol NP-10 was obtained from J. T. Baker Inc., and used without further purification. Solubilization Tests. Phenanthrene, anthracene, and pyrene were used in batch tests with 50-mL Pyrex centrifuge tubes filled to zero headspace, sealed with Teflon-lined septa, and secured with open-port screw caps.
Each tube initially received 6 g of soil spiked with a measured volume of PAH stock solution in methanol containing between 0.3 and 0.5 pCi of [14C]PAH,corresponding to -6 x 105-106dpm. The mass of PAH used in each tube was approximately 0.1, 0.8, and 2.7 mg, respectively, for anthracene, pyrene, and phenanthrene. The methanol was allowed to evaporate for 2-3 h before use. The surfactant solutions were prepared in BOD dilution water (33)and their concentrations are reported as volume of surfactant/volume of water. The sealed centrifuge tubes were mounted on a tube rotator and rotated for 15 min every 0.5 h to maintain the soil in suspension for an equilibration period of 24 h or more (32). Comparative studies with 200 mg/L HgClz as microbial inhibitor showed that surfactant biodegradation was insignificant over an initial 24-h equilibration period. Prior to analysis the tubes were centrifuged for 30 min a t 450g and aliquots withdrawn with a syringe and expressed through a preconditioned 0.22-pm PTFE membrane filter. Preconditioning consisted of rinsing with 1 mL of methanol and then purging with several milliliters of sample. Triplicate samples of 0.5-mL aliquots were injected into 20-mL polyethylene counting vials with 10 mL of liquid scintillation cocktail (Scintiverse 11, Fisher Scientific, Pittsburgh, PA) and counted for 14C. Mineralization Tests. The biodegradation tests employed an aerobic system consisting of a closed biometer flask fitted with a side tube containing sodium hydroxide solution (34,35).The system was sealed with Neoprene stoppers on both the flask and side arm. A sampling needle extended into the side arm for periodic extraction of NaOH. Each flask received 10 g of soil spiked with 4.5 mg of phenanthrene and 80 mL of aqueous solution, for a soil-to-water ratio similar to that used in the solubilization tests. The activity of [14C]phenanthrene in individual tests was between 0.3 and 0.5 pCi. The soil was kept in suspension by mounting the biometer flasks on ganged electromagnetic stir plates (15-point stirrer, Cole Palmer Instruments, Chicago, IL) with timers set to stir gently for 7 min every 0.5 h. This ensured that the soil would be suspended periodically while avoiding warming of the samples by continuous operation of the electromagnetic stir plates. The pH of the soil-water suspension at the start of the experiment was between 5 and 6 without adjustment. The final pH of the systems at the end of the tests were measured to be in the range pH 6-7. As in the solubilization tests, surfactant solutions were prepared in BOD dilution water and their concentrations are reported as volume of surfactant/volume of water. Earlier biometer studies using only the soil inoculum indicated no phenanthrene mineralization in these soilwater systems over a period of 12 weeks. Subsequent mineralization tests, therefore, were inoculated with a mixed population of PAH-degrading bacteria (RET-PA101) that had been isolated from contaminated wastes and soils. The soil used in the mineralization tests was not sterilized prior to inoculation since the indigenous bacteria appeared unable to mineralize phenanthrene over the time course of the experiments. For the purpose of inoculation, the PAH-degrading culture was grown in an autoclaved growth medium containing 100 ppm naphthalene and 1
Table 11. Structures and Properties of Selected Nonionic Surfactants surfactant Brij 30 Tergitol NP-10 Triton X-100
structure
symbol
MW
C12H26(OCH2CH2)40H dodecylethoxylate with 4 ethoxylate units C12E4 363 CgHlgC6H40(CH2CH20),Hnonylphenylethoxylate with average X = 10.5 CgPE10.6 683 CsHl7C6H40(CHzCHz0),Hoctylphenylethoxylate with average X = 9.5 CsPEg,5 625
cmc, M (vol % ) 2.3 X 10" (8.3 X lo-') 5.4 X lo6 (3.7 X 1.7 X lo-' (1.1 X
Environ. Scl. Technol., Vol. 25, No. 11, 1991
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g/L glucose to facilitate rapid bacterial growth (36). Enumeration of PAH-degrading organisms was performed by plating 0.1-mL aliquots from appropriate dilutions of the culture on phenanthrene-spread plates. The plates were prepared by spreading 0.2-mL portions of a phenanthrene-acetone mixture (5 g/L phenanthrene) on a prepared medium of mineral salts and agar (10, 36, 37). The acetone was allowed to evaporate overnight, and the inoculated plates were incubated at room temperature for 5 days or longer. The PAH-degrading colonies were identified as clear circular zones against a cloudy field. A colony counter with a UV light was employed for counting purposes. Degradation of 14C-labeledphenanthrene was detected by trapping and analyzing liberated 14C02. The side arm of the biometer flask contained 20 mL of 2 M sodium hydroxide that trapped C02. NaOH (0.5 mL) was withdrawn in triplicate for each sampling interval, placed in scintillation vials containing 10 mL of scintillation cocktail (OptiFluor, Packard Instrument Co., Downers Grove, IL), and counted for 14Cafter being stored overnight in the dark to minimize chemiluminescence. The NaOH solution was replaced after four sampling periods, during which time oxygen was contacted with the systems. Additional biometer studies in which the soil-aqueous surfactant solutioons were contacted with oxygen twice a week gave similar [I4C]phenanthrenemineralization behavior, indicating that oxygen availability per se was not a factor in the biodegradation of phenanthrene. Blanks were set up without soil and bacteria to assess for any potential volatilization of phenanthrene; controls with sterilized soil were also used to confirm results from the blanks. Calculations on cumulative mineralization employed average dpm values for each triplicate set of samples and accounted for the dpm removed a t each stage. Acidification of the biometer liquor a t the conclusion of various experiments in representative tests with more than 40 biometers indicated that 14C02release after acidification was not significant, and hence it was not necessary to account for [14C]carbonate in samples. Tests Incorporating Dilution of the Surfactant Solution. (a) Soil-Water System. Biometers were set up with nonionic surfactants a t concentrations that had been observed to inhibit microbial mineralization of phenanthrene in soil-water systems. After several weeks, the contents of the biometer flasks were centrifuged at 500g for 0.5 h. The solids were separated and resuspended in the biometer flask with fresh BOD dilution water, and an appropriate amount of supernate, to decrease the surfactant concentration to a level observed not to inhibit the mineralization of phenanthrene. Percent mineralization reported following the dilution is based on the [14C]phenanthrene remaining in the biometer after the radioactivity in the discarded supernate was accounted for. (b) In the Absence of Soil. Biometers without soil were set up to observe the mineralization of [14C]phenanthrene in systems receiving nonionic surfactants at doses of 3-5X cmc. Table I1 lists measured cmc values of the surfactants in clean water as determined from surface tension measurements (38). The contents of these systems were diluted to a surfactant concentration of 0.5X cmc after 5 weeks. This dilution involved no removal of phenanthrene or microorganisms, and mineralization is reported as the cumulative fraction [14C]phenanthrene mineralized from the start of the experiment. Mass Balance. Following mineralization of [ 14C]phenanthrene or [14C]glucose,the contents of the biometer flask were analyzed for residual 14C before disposal, to
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Environ. Sci. Technol., Vol. 25, No. 11, 1991
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1 Phenanthrene Soil:Water ratio of 1 :8
80
-
60
-
40 -
20
-
0 0 0
Y 0 5
1 0
1 5
Surfactant Dose, vot. surf./vol. water, Figure 1. Solubilizationof phenanthreneby nonionic surfactants in a soil-water system.
ascertain that mass balance constraints were satisfied. Combustion analysis of the soil-water slurry was performed in an R. J. Harvey biological material oxidizer (BMO). The soil-water mixture was combusted in the BMO and 14C02from combustion was captured directly in a scintillation cocktail containing carbamate (R. J. Harvey). The cocktail was then analyzed for 14C. Attempts to extract the PAHs from the soil with hexane, methanol, or methylene chloride by sonicating the counting the extract resulted in very poor recovery efficiencies, except in the presence of significant surfactant concentration. 14C-materialbalances by combustion analysis for a variety of biometer systems showed satisfactory recovery efficiencies ranging from 82 to 97%, with an average recovery efficiency of -91 %.
Results Surfactant Solubilization of PAH Compound from Soil. The three surface-active agents used in this study were nonionic polyethoxylate surfactants: Brij 30, C12E4, a dodecylethoxylate with 4 ethoxylate units; Triton X-100, C8PE9.5,an octylphenylethoxylate with 9-10 ethoxylate groups; and Tergitol NP-10, C9PE10.5,a nonylphenylethoxylate with an average of 10-11 ethoxylate groups. Surfactant or micellar solubilization is the dissolution of a substance by reversible interaction with the micelles of a surfactant in a solvent to form a thermodynamically stable isotropic solution (39). Solubilization is initiated at a surfactant concentration known as the critical micelle concentration (crnc), where surfactant molecules aggregate to form micelles. Figure 1 shows the solubilization (Le., mass fraction in the liquid phase) of phenanthrene by the C12E4and CgPE9.6e surfactants, in a soil-water system with a soil to water ratio of 1:8 (w/v). Data in Figure 1show that phenanthrene solubilization in the soil-water system is observed only at surfactant doses >0.1%, which is many times greater than the cmc, i.e., -8.3 X (v/v) for (v/v) for C8PE9.5.Additional tests C12E4 and 1.1 x conducted with surfactant concentrations of CO.l% confirmed that at lower surfactant doses solubilization was not significantly different from the aqueous solubility of the PAH compound. The presence of soil changes the surfactant micellization pattern due to surfactant sorption onto the soil, resulting in the aqueous-phase surfactant being considerably less than the total added. Subsequent surface tension experiments with soil-water systems confirmed that in the presence of soil, surfactant doses necessary to attain the cmc were considerably greater than in clean water systems. The surfactant doses at which micelles form in the presence of 6.25 g of soil/50 mL total
-
OU
Phenanthrene Sokwater = 1 :8 g/mL
Phenanthrene With soil
I
without surfactant,
Soii-water system without inocuium
Control n
0 0
2
4
, IL S 8
6
10
Time, weeks Flgure 2. Microbial minerallzation of phenanthrene in a soil-water system without surfactant showlng 60% minerallzation in 10 weeks.
volume are in the range of 0.1% (38). Thus, the surfactant dose that initiates PAH solubilization from soil corresponds to that dose necessary to attain aqueous-phase surfactant micelles, following surfactant sorption, at equilibrium with maximum surfactant monomer concentration, i.e., surfactant monomer concentration equal to the cmc. Vigon and Rubin (40)observed surfactant sorption on soil for a nonylphenylethoxylate nonionic surfactant with eight to nine ethoxylate groups, and two dodecylphenylethoxylates with five and eight to nine ethoxylate units, using a prepared soil with 4% organic carbon and a soil to water ratio of approximately 1 g/100 mL. At a surfactant dose of 0.1%, 65-95% of the surfactant was sorbed onto the soil. The sorption of C12E4, CaPEe,5,and C9PE10.5nonionic surfactants onto soil has been reported by Liu et al. (38). Data as in Figure 1, showing the mass fraction of total PAH in the liquid phase based on [I4C]PAHanalysis, have been obtained for pyrene and anthracene in similar soilwater systems with total PAH loading adjusted to attain aqueous-phase solubility concentration in the absence of surfactant (32). The solubilization experiments suggest that a surfactant concentration of 1% (v/v) with a soil to water ratio of 123 (g/mL) would transfer substantial amounts of the PAH compound from the soil-sorbed phase to the liquid phase of the aqueous surfactant solution. Accordingly, microbial mineralization experiments were performed with surfactant concentrations ranging from 0 to 1%. Mineralization of [ 14C]Phenanthrenei n the SoilWater System. Biological mineralization of phenanthrene was monitored by the capture of 14C02in caustic solution from a soil-water system comprising 10 g of soil, 80 mL of BOD water (33),and 4.5 mg of phenanthrene, thereby maintaining conditions similar to those in the solubilization tests. Each biomineralization test was performed in duplicate, as were tests comprising blanks and sterile controls. Each biometer received a 2-mL inoculum of PAH-degrading culture grown earlier on a mineral medium with naphthalene and glucose. The PAH-degrading culture contained an average of 1.3 X lo7colony-forming units/mL (CFU/mL) when assayed by tube dilution and spreading on phenanthrene plates. [14C]Phenanthrenelosses through volatilization or abiotic degradation were assessed by the use of autoclaved soil controls. Figure 2 shows the mineralization of [*4C]phenanthrenein the soil-water system in the absence of surfactant. Over the course of 10 weeks nearly 60% of the initial [14C]phenanthrenewas mineralized by the PAH-degrading culture. Without the inoculum there was no appreciable mineralization of phenan-
-
0
2
4
6
8
10
12
Time, weeks Flgure 3. Comparison of phenanthrene minerallzation in clean water and in a soil-water system. Both systems received the same PAHdegrading inoculum.
threne. The 14Canalysis for the autoclaved soil control is also shown. The absence of 14Cin the caustic solution for the control indicates that abiotic processes do not contribute to the measurement of phenanthrene mineralization. Figure 3 shows the results of experiments comparing the microbial mineralization of phenanthrene in another soil-water system with a similar system receiving no soil and having the same PAH-degrading inoculum and mass of phenanthrene. In the absence of soil, the phenanthrene loading greatly exceeds aqueous solubility, and thus crystalline phenanthrene is present. As in Figure 2, for the soil-water system there is a short lag interval followed by a period of relatively rapid I4CO2production. In the absence of soil, no significant evolution of '%02is observed until about the fourth week, and the subsequent 14C02 production rate is less. The extent of mineralization for the soil-water system is much greater than that observed for the system without soil over the course of 12 weeks. A possible explanation for this observation is the greater availability of PAHs from the soil than from the dissolution of solid phenanthrene; Le., in the absence of soil, microbial mineralization may be limited by dissolution of the crystalline substrate. Mineralization of ['%]Phenanthrene i n the Presence of Surfactant. The mineralization of [I4C]phenanthrene by PAH-degrading organisms in soil-water systems was examined at varying concentrations of the nonionic surfactants. Although a 1% surfactant dose effectively solubilized PAH from the soil phase to the liquid, the presence of any of the three surfactants at 1% concentration completely inhibited the mineralization of phenanthrene. The results for the biometers receiving the C12E4 surfactant are shown in Figure 4. The rate of mineralization of [I4C]phenanthrene is maximum in the absence of surfactant, proceeding almost at the same rate in the presence of 0.01% (v/v) C12E4. However, there was no significant mineralization observed in the soil-water systems for C12E4 surfactant doses of 0.05% (v/v) and greater. Phenanthrene mineralization in the presence of C,PQ,5 is shown in Figure 5. The microbial mineralization of phenanthrene is inhibited at surfactant concentrations in excess of 0.05% (v/v). Figure 6 shows that the mineralization of phenanthrene was completely inhibited in soil-water systems in the presence of 0.2% (v/v) CgPEl0.,. This concentration is somewhat greater than the surfactant dose found to be Environ. Scl. Technol., Vol. 25, No. 11, 1991
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80 i Phenanthrene Soikwater = 1:8 g/mL Surfactant Brii 30, C12E4
Phenanthrene
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-0- 0.20%
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Time, weeks Flgure 4. Inhibition of phenanthrene biodegradation in a soil-water system receiving various doses of &E4 nonionic surfactant. Surfactant doses of 0.05% (v/v) and greater completely arrested microbial mineralization. on
Phenanthrene 60 -
0
2
4
a
6
Time, weeks Flgure 6. Inhibition of phenanthrene biodegradation with CBPE, nonionic surfactant at a surfactant dose of 0.2% (v/v). Lower surfactant concentrations did not enhance the rate or extent of mineralization of phenanthrene.
Soil:water = 1.8 g/mL Surfactant Triton X-1 00, C8PE9.5
Giucose Soil:water = 1 :8 g/mL
8 i
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20-
0.10% 0.20%
V
% 0
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. Glucose Soil:water = 1 :8 g/mL
8
Time, weeks Flgure 5. Inhibition of phenanthrene biodegradation in a soil-water system receiving various doses of C8PE,,, nonionic surfactant. Inhibitory effect was observed at surfactant concentrations greater than 0.05% (v/v).
limiting for the other two surfactants tested. As discussed later, the surfactant dose above which microbial mineralization of phenanthrene is inhibited appears related to the cmc of the surfactant in the presence of soil. The inhibitory effect of higher surfactant doses on phenanthrene mineralization may be due to various phenomena, including toxic effects, the preferential use of surfactant as substrate, the lowering of aqueous-phase PAH concentration due to micellization, or an interference of the surfactant with microbial metabolic processes. Subsequent experiments evaluated some of these possible causes. Glucose vs Phenanthrene Mineralization in the Presence of Surfactant. The inhibition of microbial mineralization a t 1%(v/v) surfactant dose might be due to toxicity of the surfactant at higher concentrations or due to toxicity of the solubilized phenanthrene present a t much higher concentrations in the aqueous surfactant phase. To evaluate surfactant and micellized-phenanthrene toxicity to the microorganisms, [14C]glucosemineralization in the soil-water system was investigated by using the following biometer configurations with PAHdegrading cultures: (i) [14C]glucosealone, (ii) [14C]glucose and phenanthrene, (iii) [14C]glucosewith 1%CI2E4,and (iv) [14C]glucoseand phenanthrene with 1%C12E4.Results in Figure 7A show that there was no effect of the presence 1924 Environ. Sci. Technol., Vol. 25, No. 11, 1991
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Time, days
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Giucose Soil water = 1 8 qimL
Giucose alone
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-
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-c
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with 4.5 mg phenanthrene and 1% Brij 30
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C .
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.
.
.
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.
.
.
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Time, days Flgure 7. Mineralization of radiolabeled glucose in soil-water systems. Degradation was not affected by the presence of phenanthrene (A) or by 1% (v/v) without phenanthrene (B) or with phenanthrene
(C).
of phenanthrene on glucose mineralization. C12E4 surfactant at 1%(v/v) (Figure 7B), or phenanthrene together with 1% C12E4surfactant (Figure 7C), did not significantly affect the mineralization of [14C]glucosein the soil-water system. Thus, neither micellar surfactant nor micellized
Phenanthrene
Glucose
No soil
No soil
8
8-
1
c
.-0 m .-N E
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a, C
I
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20
30
40
Time, days
Time, weeks
Figure 8. Mineraiization of [ ''C]glucose in clean water. Degradatlon was not signlficantly inhibited in the presence of nonionic surfactants at doses greater than the cmc.
PAH was inhibitory to the glucose-degrading microorganisms in soil-water systems. Additional tests were performed without soil in order to evaluate the specific effect of surfactant on the PAHdegrading inoculum in the absence of the mixed consortium of organisms associated with the soil. These tests ascertained the effect of the surfactant on [14C]glucose mineralization with 0.2% (v/v) of individual surfactants, corresponding to surfactant doses that ranged from 1OX cmc (e.g., C8PE9,5)t o -240X cmc (e.g., C12E4). Figure 8 shows that the mineralization of [14C]glucose by the PAH-degrading inoculum was not impaired by surfactants present at a concentration of 0.2% (v/v). Since a mixed population of unknown bacteria capable of degrading PAH compounds was used in these tests, it is likely the inoculum may contain glucose-degradingstrains with very different sensitivities to the presence of surfactant as compared with the PAH-degradingstrains. Nonetheless, results in Figures 7 and 8 suggest that microbial toxicity by the surfactant may not be the specific cause for the inhibition of biomineralization of phenanthrene. A parallel set of experiments was performed to evaluate phenanthrene mineralization in systems without soil. The data from these tests are presented in Figure 9 and show that, in the presence of 0.2% (v/v) of any of the three nonionic surfactants considered, microbial mineralization of [ 14C]phenanthreneis completely inhibited in the systems recieving no soil. Also, in the absence of both soil and surfactant, mineralization of [14C]phenanthreneproceeds much slower and to a lesser degree compared to a soil-water system, confirming results shown in Figure 3 (Le., less than 10% mineralization over a period of 8 weeks for a no-soil system compared to almost 60% for the soil-water system). The growth of pure Escherichia coli culture on glucose was studied in the presence and absence of two nonionic surfactants. Bacterial cell density was monitored by measuring the absorbance at 560 nm, and the specific growth rates observed in the presence of 0.1?' & (v/v) Triton X-100 (octylphenyl(9.5)ethoxylate surfactant) or Tween 80 (sorbitan monooleate(20)ethoxylate surfactant) were similar to the specific growth rate observed in the absence of surfactant. This indicates that the surfactants at the supra-cmc concentration did not inhibit growth on glucose of the pure Gram-negative E. coli strain; this is a further confirmation that the surfactants did not produce an inherent toxic effect.
Flgure 9. Mineraiization of [ "Clphenanthrene in clean water. Degradation was completely inhibited In the presence of nonbnlc surfactant at doses greater than the cmc.
-0-
No surlactant 0.05% Triton X-100 0.05% Brij 30
Phenanthrene Soil:water = 1.8 g/mL Initial surfactant concentration of 0.05% v/v to -0.01% at week 5
-
0
2
4
6
8
10
12
Time, weeks Flgure 10. Mineraiiration of phenanthrene in soil-water systems. Degradation commenced following dliution of nonionic surfactant from 0.05 (v/v) to -0.01 % (v/v).
Effect of Dilution on Phenanthrene Mineralization. To further corroborate that surfactant toxicity to the microorganisms per se is not a factor responsible for the inhibition of phenanthrene mineralization at higher surfactant doses, the effect on dilution of the surfactant solution was considered. Figure 10 shows the recovery of phenanthrene mineralization observed in soil-water systems when the surfactant was diluted. The original surfactant solution was 0.05% (v/v) for both the CI2E4and C8PE9,5.After 5 weeks there was no significant mineralization of [ 14C]phenanthreneobserved in the presence of 0.05% (v/v) surfactant, whereas in the absence of surfactant, mineralization in the soil-water system exceeded 40%. These data at 5 weeks are consistent with results shown in Figures 4 and 5. After 5 weeks, the contents of the biometer flasks were centrifuged, and the aqueous surfactant phase was decanted and replaced with BOD dilution water. Following dilution of the surfactant to a dose of -0.01% (v/v), the system recovers bioactivity with respect to phenanthrene mineralization such that, 5 weeks after dilution, mineralization of the remaining PAH is 30 and 20% for C8PEg.5 and C12E4, respectively. The mineralization of phenanthrene in the absence of soil was evaluated with dilution of surfactant solution. Environ. Sci. Technol., Voi. 25, No. 11, 1991
1925
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No surfactant ---C
50
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0.072% Triton X-100
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Phenanthrene Soi1:water = 1:8 gimL with glucose-glutamic acid
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Phenanthrene 60
No soil, surfactant initially present at 0.072% viv, diluted to -0.5 x CMC at week 5
40
30 -
glucose-glutamic acid mixture, mg
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20 - 1
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* 10
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-A-
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2
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Time, weeks Figure 11. Minerallzation of phenanthrene in the absence of soil. Degradation commenced following dilution of nonionic surfactant to a concentration less than cmc.
Figure 11 shows the mineralization of phenanthrene for a system receiving no soil or surfactant, and another system receiving no soil and an initial concentration of 7.2 X lo-*% C8PE9,5,Le., 7X cmc as evaluated from surface tension measurements in clean water (38). Both systems exhibited no phenanthrene mineralization in the first 5 weeks. After 5 weeks, the surfactant in the system receiving 7X cmc surfactant concentration was diluted to 0.5X cmc by the addition of BOD dilution water. Following dilution, mineralization proceeded without a significant lag period and reached a level of 30% after 7 weeks for the system now a t 0.5X cmc. Phenanthrene mineralization was less than 20% in a 12-week period for the system without surfactant. The presence of a small amount of surfactant (-0.5X cmc) in the system without soil appears to enhance the biomineralization of phenanthrene. This observation agrees with earlier reports (13, 24, 26, 41) that dilute surfactant solutions increased the rate of mineralization of sparingly soluble, separate phase, hydrophobic compounds in aqueous systems by rendering them more bioavailable. Phenanthrene Degradation in the Presence of Glucose. Phenanthrene degradation in the presence of glucose was studied to assess the effect of the presence of a preferred substrate. Figure 12 shows the mineralization of phenanthrene in soil-water systems receiving from 0 to 100 mg of a 5050 glucose-glutamic acid mixture/80 mL water. The presence of 1 or 10 mg of glucose-glutamic acid/80 mL of water (corresponding to 13 and 130 ppm, or 0.0013 and 0.013% (w/v), respectively) has no effect on phenanthrene mineralization over a 5-week period. The addition of 100 mg of glucose-glutamic acid/80 mL of water (Le., 0.13% w/v) resulted in a short lag interval followed by rapid evolution of 14C02. These data indicate that the presence of glucose as a preferred substrate a t levels as high as 0.13% (w/v) did not have a significant inhibitory effect on the degradation of phenanthrene.
Discussion The data show that phenanthrene mineralization in a soil-water system inoculated with PAH-degrading organisms proceeds at a rapid initial rate, declining after -8 weeks with approximately 60% of the phenanthrene being mineralized to COz. The most significant effect of the addition of higher concentrations of nonionic ethoxylate surfactants to the soil-water systems was the essentially 1026
Environ. Sci. Technoi., Vol. 25, No. 11, 1991
Time, weeks Flgure 12. Phenanthrene mineraiization in soil-water systems. The presence of increasing concentrations of giucose-glutamic acid mixture was not inhibitory.
complete inhibition of phenanthrene mineralization. The inhibitory effect of surfactant on biodegradation may be a result of various possible influences on microbial degradation including the following: (i) a toxic effect of the surfactant, (ii) a toxic effect of the micellized PAH, (iii) a reduction in the equilibrium free aqueous-phase concentration of phenanthrene to less than a threshold value, (iv) a preferential use of surfactant as substrate in lieu of PAH, (v) a physicochemical effect of micelles interfering with microbial enzymatic activity or with the transport of PAH to and/or across the microbial cell envelope, or (vi) bioavailability of micellized PAH. Some of these possible effects are discussed below. Apparent Nontoxicity of Surfactant and/or Micellized Phenanthrene. The inhibitory effect of the nonionic surfactants in soil-water systems was evident a t surfactant doses greater than a range of about 0.05-0.1% (v/v) for soil to water ratios of 123 (g of soil/mL of water). As discussed later, this concentration range corresponds to that which results in the formation of micelles in the aqueous phase, causing desorption and solubilization of PAH compounds via micellization. Phenanthrene mineralization was observed in surfactant systems upon dilution to sub-cmc values for either soil-water or aqueous solutions. These results suggest that it is the presence of surfactant micelles, not the surfactant per se, that inhibit microbial mineralization of phenanthrene. Glucose was mineralized readily in soil-water systems with 1% (v/v) surfactant, and in clean water systems with 0.2% (v/v) surfactant corresponding to about 20-240 times the cmc. Glucose mineralization was not affected by the presence of either sorbed phenanthrene or micellized phenanthrene in soil-water systems with 1% CI2E4.The results with glucose suggest that neither the surfactant nor the micellized phenanthrene is inherently toxic to the microorganisms. It is important to note, however, that the sensitivity to surfactant of the PAH-degrading strains may be very different from that of glucose-degrading strains. Nonetheless, the recovery of phenanthrene mineralization upon dilution of soil-water systems with surfactant suggests an apparent nontoxicity of the nonionic surfactants at the doses used. Thus, the data suggest that the influence of the nonionic surfactant on inhibition of phenanthrene mineralization is related to a physicochemical effect rather than a specific toxic effect. Micellization and Free Water Concentration of Phenanthrene. The process of micellization of phenan-
Phenanthrene Solubility at Aq. CMC
61
n ., ,
0.0
.
I
0.5
Surfactant
1 .o
Dose, O/O
1.5
2.0
(V/V)
Figure 13. Predicted free aqueous-phase concentration of phenanthrene in a soil-water system receiving Increased dose of CBPEB5 nonionic surfactant. The model parameters were Ob",the mass of surfactant sorbed per unit mass of soil, 1.9 X 10- moi/g; K,,,,, soli-aqueous-phase partition coefficient for phenanthrene in the presence of maximum surfactant monomer concentration (Le., cmc), 327 mL/g; and K,, the phenanthrene mole fraction micelle-aqueous-phase partition coefficient, given by log K , = 5.7 (57).
threne in soil-water systems results in solubilization of sorbed-phase phenanthrene to the aqueous phase via incorporation in surfactant micelles. The microbial degradation of phenanthrene would be affected if the free water concentration of phenanthrene was diminished substantially as a result of partitioning to surfactant micelles. The following describes a modeling approach to assess whether surfactant solubilization of phenanthrene affects the free aqueous-phasephenanthrene concentration. The modeling approach (57) is based on a series of companion investigations that describe PAH solubilization (e.g., Figure 1) in the presence of nonionic surfactant considering the following: (i) surfactant micelle-PAH partitioning, (ii) surfactant sorption onto soil, (iii) effect of surfactant sorption on soil organic carbon content and increase in PAH-soil sorption coefficient, and (iv) effect of surfactant monomers at a concentration equal to the cmc on PAH solubility. The solubilization of hydrophobic organic compounds (HOCs) by nonionic surfactant micelles in a soil-water system entails the simultaneous transfer of the organic compound from the soil-sorbed phase to the aqueous phase, and from the aqueous phase to the hydrophobic interior of the nonionic surfactant micelles. In the absence of excess hydrophobic organic compound, the reservoir of HOC available for micellar solubilization is limited to the HOC in the sorbed phase and in the aqueous phase. The solubilization process accordingly tends to deplete the amount of HOC in each of these nonmicellar compartments. The concentration of HOC in the aqueous phase, denoted here as C,,, varies as a function of the mass of surfactant added to the system. C,, can be estimated for a system of soil and micellar surfactant solution from a relationship based on HOC mole balance for a soil-water system before and after surfactant addition (57). As shown in Figure 13 for a soil-water system with 10 g of soil, 80 mL of water, and 4.5 mg of phenanthrene, Caq,the concentration of phenanthrene in the aqueous phase, is a monotonically decreasing function of the dose of the surfactant C8PE9.& added. At a surfactant dose of 1% (v/v), the aqueous-phase concentration of phenanthrene is 27% of phenanthrene solubility. This corresponds to a concentration of -0.36 mg/L, which does not fall below a threshold level that might be expected for biodegrada-
-
tion, since phenanthrene biodegradation has been observed in natural systems at nanogram and microgram per liter concentrations (7). For example, Sherill and Sayler (42) showed that the rate of biodegradation of phenanthrene in natural aquatic systems was related to the phenanthrene concentration; decreasing the concentration 100-fold, from 1000 to 10 bg/L, resulted in only a 3-fold decrease in the degradation rate. Additional modeling to simulate conditions during degradation has shown that there is a relatively proportional effect of micellization on free aqueous-phase phenanthrene concentration as the total mass of phenanthrene is depleted. Thus, the relatively modest reduction in free aqueous-phase concentration of phenanthrene in the presence of surfactant micelles cannot explain the inhibition of phenanthrene mineralization observed. Degradation of Crystalline Phenanthrene. The data in Figures 3 and 11 show that the rate of mineralization is significantly less for crystalline phenanthrene in clean water than for mineralization of phenanthrene in soilwater suspension. The effect of the presence of a sub-cmc concentration of surfactant C8PE9.5(Figure 11)shows that the mineralization rate of crystalline PAH is slightly enhanced with surfactant at doses not exceeding the aqueous cmc in comparison to that for clean water. This is probably a consequence of the increased rate of dissolution of crystalline PAH due to surfactant influence an wetting and surface tension. Nonionic surfactants at doses that approach the cmc may increase phenanthrene solubility slightly in clean water, e.g., 30-40'31 (27), and this also may assist the rate of biodegradation in clean water systems. The process of hydrocarbon uptake by microbial cells occurs predominantly as solubilized or accommodated substrate (18). The first step in hydrocarbon degradation is the introduction of molecular oxygen into the solubilized or accommodatedhydrocarbon via cell-associated enzymes (43), implying the rate of solubilization or dissolution of PAHs may control growth. This is consistent with the conclusions of Wodzinski and Coyle (44) that the phenanthrene mineralization rate is limited by solubility and is supported by calculations based on growth rates of bacteria on crystalline phenanthrene, for which the dissolution rate limits the rate of degradation (45). Similar inference is evident in the data of Thomas et al. (46) for naphthalene and 4-chlorobiphenyl. Surfactants may act to emulsify liquid hydrocarbons and to improve mass transfer of solid hydrocarbons, thereby increasing their availability to microorganisms (26). The significance of the role of dissolution and emulsification is borne out by several studies involving the use of surfactant solutions in liquid culture media to enhance the rate of microbial degradation of insoluble hydrocarbon substrates. Breuil and Kushner (23) investigated the effects of the surfactants C8PEg,,(Triton X-100) and C12E23 (Brij 35) on the bacterial utilization of hexadecane. The surfactants were not themselves used for growth, but they stimulated bacterial growth on hexadecane and greatly decreased the lag period before hexadecane utilization at concentrations of 0.001-0.01 %. These doses are less than the cmc concentrations (cmc values for C8PE9.5and CI2E, are about 0.011 and 0.07% (v/v), respectively). Guerin and Jones (26) reported that the use of various Tween-type nonionic surfactants in aqueous media solubilized phenanthrene to different degrees and enhanced phenanthrene utilization. However, the order of enhancement did not correlate perfectly with increased solubility, suggesting physiological as well as physicochemical effects of the surfactants (26). Envlron. Sci. Technol., Voi. 25, No. 11, 1991
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Preferential Substrate Utilization. A possible explanation for the inhibition of microbial mineralization of phenanthrene might be the preferential biodegradation of the nonionic surfactants. Figure 12 suggests that the presence of a readily degradable substrate, such as glucose, a t higher concentrations may result in a lag period occurring before the initiation of phenanthrene mineralization. However, even in the presence of 100 mg of glucose-glutamic acid mixture in the soil-water slurry (corresponding to -0.13% w/v), phenanthrene mineralization had reached more than 20% by the end of the second week. In addition to being readily degradable, glucose is also a highly polar organic compound and, therefore, will neither be significantly affected by the presence of micellar surfactants nor sorb on to soil, but will be present entirely in the aqueous phase. The nonionic surfactants at concentrations in the range of 0.05-0.190 (v/v) and higher were observed to inhibit phenanthrene mineralization. These concentrations correspond to surfactant loadings of about 40-80 mg in the soil-water system, which is a smaller organic loading than that for the case of 100 mg of glucose-glutamic acid mixture. The ethoxylate surfactants are not as easily metabolized as substrates such as glucose, and surfactant sorption onto soil may result in a depletion of bioavailable surfactant. These considerations contraindicate the likelihood of the surfactant being used as a preferred substrate in lieu of phenanthrene, although further growth experiments are required to verify this conclusion. Most surfactant biodegradation studies have employed surfactant concentrations much lower than those considered in the current experiments since, for the main part, prior work dealt with the environmental fate of surfactants in wastewaters. Biodegradation of nonionic ethoxylate surfactants is speeded by increased linearity of the hydrophobe and retarded by increased degree of ethoxylation, i.e., length of the E chain (48). The linear alcohol ethoxylates are characteristically biodegradable. Using samples of C16E3and CI2Eglabeled either in the hydrophobe or in the E chain, Larson and Games (49) studied the rate and extent of 14C02formation in river water and a mineral medium. At initial surfactant concentrations of 0.001-0.1 ppm, mineralization was observed immediately, whereas there was a 2-10-day lag before mineralization of surfactant was observed when surfactant concentrations were increased to 0.5-10 ppm. The concentration of C12E4used in the phenanthrene mineralization tests was -0.1 %, which corresponds to 1000 ppm. Surfactant mineralization under these circumstances may entail a still longer lag period during which time phenanthrene mineralization would be evident if unimpeded. The present study employed two alkylphenol ethoxylate surfactants: C8PE9,5and CgPE10,5.These surfactants are a subset of those selected earlier for screening solubilization tests with PAH compounds in soil-water experiments (32). In possible environmental application, the alkylethoxylates would be preferred over the alkylphenol ethoxylates owing to the undesirable degradation products of the latter (50). Laboratory experiments have shown that the biodegradation of alkylphenol polyethoxylates leads to the formation of biorefractory intermediates with one and two oxyethylene groups. Osburn (51) examined the biodedgradation of t-C8PE, and showed that the E chain was degraded in river water down to C8PE2.5.Similar observations have been reported in a number of studies involving river water, treated municipal wastewaters, digested sewage sludges, and tap water (52-55). The tox1928 Environ. Sci. Technol., Vol. 25, No. 11, 1991
icities for alkylphenol ethoxylates generally increase with shorter oxyethylene side chains, suggesting that the alkylphenol polyethoxylate surfactants may be biologically transformed to refractory metabolites that are more toxic than the original chemicals. In summary, although nonionic surfactants are susceptible to microbial degradation to some extent, these surfactants are not as easily degraded as glucose. Further, in consideration of surfactant sorption onto soil removing a significant mass fraction, it may be reasonable to conclude that the inhibition of phenanthrene degradation observed at surfactant doses that resulted in micelle formation is not a result of the preferential substrate utilization of the surfactant by the PAH-degrading microorganisms in lieu of phenanthrene. Surfactant-Bacteria Interactions. Surfactant toxicity or inhibition may result from (i) disruption of cellular membranes by surfactant interaction with their lipid structural components and (ii) reaction of surfactant with the enzymes and other proteins essential to the proper functioning of the bacterial cell. Nonionic surfactants are in general less active against bacteria (48). Bacteriotoxicity tends to increase with increased hydrophobic chain length and, in ethoxylate nonionics, to decrease with increased E-chain length. Cell membranes are comprised, a t least in part, of a bilayer structure made of phospholipid-type surfactants. These microbial surfactants can form membranous micellar double layers, arranging themselves so that the two hydrophilic groups form two parallel boundaries with the hydrophobic tails aligned to fill the space between. The phospholipid bilayer isolates aqueous cell or organelle interior from the outside by virtue of the hydrophobic inner layer. Foreign surfactant molecules may disrupt cellular membranes by substituting for, or otherwise interfering with, the phospholipids, causing loss of microbial cell contents to the exterior either by penetration of cell membrane or mechanically upon its rupture (48). The formation of complexes between surfactants and proteins constitutes a reaction of surfactant that may lead to changes in shape and activity of enzyme protein and to weakening or dissolution of structural protein. Nonionic surfactants form complexes with membrane proteins that are more or less fixed in the cell membrane (48). Formation of surfactant-protein complexes makes it possible to disengage the membrane proteins from the membranes. Membrane proteins are believed to play a significant role in the transport of materials across the cell wall, and the formation of surfactant-protein complexes may interfere with the usual transport mechanisms. Such binding with membrane proteins occurs at nonionic surfactant concentrations near or above the cmc, perhaps by direct interaction with preformed micelles (48). However, even though nonionic surfactants may displace the membrane lipids from the hydrophobic areas of the protein, they are not as powerful as ionic surfactants in altering the protein conformation and generally result in a reversible alteration (48). These considerations may explain the inhibition observed in the microbial utilization of phenanthrene in the presence of surfactant at concentrations above the cmc, and the apparent reversal of such inhibition upon dilution. Various studies indicate the absence of the involvement of extracellular enzymes for the microbial degradation of hydrocarbons (43). This would imply that transport into the cell is a prerequisite for biodegradation to occur and supports a belief that supra-cmc surfactant inhibition is a transport-related phenomenon. Miller and Bartha (56) indentified transport limitation as the cause for CM alkane
recalcitrance. They performed tests using small unilamellar vesicles called liposomes to facilitate transport through the microbial cell wall, delivering the encapsulated substrate directly to the membrane-bound enzyme systems of the intact cell. I t may be that the phenanthrene-degrading microorganisms also employ such vesicles or related structures to transport phenanthrene from the aqueous medium into the bacterial cell. Therefore, the presence of micellar surfactant may result in an interference either of the transport mechanism or of the membrane-bound protein activity. Bioavailability of Micellar PAH. The bioavailability of micellar PAH to microorganisms also needs investigation. Although the literature suggests that there is a rapid exchange of HOCs between the aqueous phase and surfactant micelles, there is only limited information on exit rates of HOCs from micelles (58,59). I t is possible that the surfactant systems considered in this study did not attain equilibrium readily in response to microbial degradation of aqueous-phase phenanthrene; in which case, exit rates for the PAH from micelles may be slow, resulting in the PAH being trapped within the micelle and unavailable for microbial degradation. This may inhibit phenanthrene biodegradation in systems with micellized surfactant.
Conclusion Many recalcitrant hydrophobic organic compounds observed in soil and sediment environments are sparingly soluble in water and strongly sorbed to soil-sediment material. It has been suggested that sorption onto inert matrices makes HOCs less bioavailable for degradation (3, 28). These hydrophobic organic compounds can be mobilized, Le., transferred from the soil-sorbed phase to the aqueous phase, by decreasing the interfacial tension between the compound and water, and thereby presumably rendered move available for microbial attack (28). Reduction of interfacial tension can be achieved by the application of surfactants, dispersants, extractants, and emulsifiers (28). The addition of surfactants to contaminated soil sites may then serve as a possible amendment in bioremediation efforts. The current study indicates that a t sub-cmc surfactant levels the degradation of phenanthrene proceeds a t best as fast as in the absence of the three nonionic surfactants employed in this study. Sub-cmc doses of surfactant in such soil-water systems do not inhibit mineralization of phenanthrene but neither do they enhance the degradation rate. At surfactant concentrations in excess of aqueousphase cmc, these surfactants exhibited an inhibitory effect on phenanthrene mineralization. This supra-cmc inhibitory effect is believed not to be attributable to a specific toxic effect per se, either of the surfactant or of the micellized PAH, nor is it thought to result from a preferential use of surfactant as substrate in lieu of phenanthrene, or due to reduction of the equilibrium, free aqueous-phase phenanthrene concentration below a threshold level that might be necessary for phenanthrene utilization. The inhibition may be a result of a physical-chemical effect of the surfactant micelles interfering with substrate transport into the cell, or with the activity of enzymes and other membrane proteins of the cell. The inhibition may also be a result of limited bioavailability of micellar phenanthrene due to low exit rates from the micelles. The three surfactants considered in this study were selected on the basis of their PAH-solubilizing capacities in soil-water systems. Additional work is necessary to assess the specific mechanism for inhibition of microbial degradation of PAH by nonionic surfactant micellization,
and whether this effect is specific to the surfactants employed in this work. Studies are in progress to address these questions by use of mixtures of nonionic surfactants, polysorbate, and other surfactants. Registry No. Brij 30,9002-92-0; Tergitol NP-10, 9016-45-9; Triton X-100,9002-93-1; phenanthrene, 85-01-8; glucose, 50-99-7.
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Surfactants onto Soil, submitted to Water Res. (39) Rosen, M. J. Surfactants and Interfacial Phenomena, 2nd ed.; John Wiley and Sons: New York, 1989. (40) Vigon, B. W.; Rubin, A. J. J.-Water Pollut. Control Fed. 1989,6I, 1233-1240. (41) Robichaux, T. J.; Myrick, H.N. J . Pet. Technol. 1972,24, 16-20. (42) Sherrill, T. W.; Sayler, G. S. Appl. Enuiron. Microbiol. 1980, 39, 172-178. (43) Rosenberg, M.; Rosenberg, E. J. Bacteriol. 1981,148,51-57. (44) Wodzinski, R. S.; Coyle, J. E. Appl. Microbiol. 1974, 27, 1081-1084. (45) Stucki, G.; Alexander, M. Appl. Environ. Microbiol. 1987, 53, 292-297. (46) Thomas, J. M.; Yordy, J. R.; Amador, J. A.; Alexander, M. Appl. Environ. Microbiol. 1986, 52, 290-296. (47) Asther, M.; Lesage, L.; Drapron, R.; Corrieu, G.; Odier, E. Appl. Microbiol. Biotechnol. 1988, 27, 393-398. (48) Swisher, R. D. Surfactant Biodegradation; Surfactant Science Series 18; Marcel Dekker: New York, 1987. (49) Larson, R. J.; Games, L. M. Environ. Sci. Technol. 1981, 15, 1488-1493.
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Received for review January 2,1991. Revised manuscript received June 12,1991. Accepted June 26,1991. This work was sponsored by the US.EPA Office of Exploratory Research under Grant R-815235-01-1.
Methylchloroform and Tetrachloroethylene in Southern California, 1987-1 990 Mohamed W. M. Hlsham and Danlel Grosjean” DGA, Inc., 4526 Telephone Road, Suite 205, Ventura, California 93003
By use of on-site electron capture gas chromatography, ambient levels of methylchloroform and tetrachloroethylene were measured in 1987-1990 at 21 southern California locations. The highest concentrations recorded were 61 ppb for CH3CCI3and 20 ppb for C2C1,, with 24-h averages of up to 28 ppb for CH3CC13and 12 ppb for C2CI4. Concentration ratios to that of a “background” site, San Nicolas Island, were 5-31 for CH3CC13and 1.3-6.2 for C2Cl,. At coastal and central locations, ambient levels during the fall were higher than during the summer, with fall summer ratios of up to 6 for both compounds. Spatial an diurnal variations are compared to those of peroxyacetyl nitrate (PAN) and are discussed in terms of emissions and transport.
/
Introduction
In the United States alone, some 2.7 billion pounds of toxic chemicals were discharged into the air in 1987 ( I ) . These toxic chemicals include a number of chlorinated hydrocarbons, which are used extensively as solvents, synthetic feedstocks, and household products and in the production of textiles and plastics. Many of these compounds are toxic and exhibit mutagenic and/or carcinogenic properties. Recent proposals to strengthen the Montreal Protocol on substances that deplete the ozone layer include the phase-out of chlorinated hydrocarbons such as methylchloroform and carbon tetrachloride (2). Information on ambient levels of chlorinated hydrocarbons 1930 Environ. Sci. Technol., Vol. 25, No. 11, 1991
is of direct interest in the context of regulatory action concerning toxic air contaminants and global air pollution. This study focuses on two important anthropogenic chlorinated hydrocarbons, methylchloroform and tetrachloroethylene. In the Northern Hemisphere, the “background” concentration of methyl chloroform was 158 ppt in 1985 and has been increasing at a rate of 8 ppt per year (3,4). Tetrachloroethylene is more reactive and as such does not have a global background concentration, but is ubiquitous in urban and industrial atmospheres. In southern California (Los Angeles urban area), some 12OOO metric tons of tetrachloroethylene and 13000 metric tons of methylchloroform are released every year (5). With these chlorinated hydrocarbons as a “signature”, transport of the Los Angeles urban plume eastward as far as southern Nevada has been documented (6,7). We have measured ambient levels of methylchloroform (CH3CC13)and tetrachloroethylene (CC12=CC1,) at 21 locations in southern California. These measurements were made as part of five field surveys carried out in 1987-1990. In 1987, measurements were made simultaneously at up to nine locations (8)as part of the Southern California Air Quality Study (9). In 1988-1990, measurements were made as part of surveys of air quality at 10 museums in southern California (10-13). In 1989, measurements were made as part of an air quality study carried out a t two locations in the eastern end of the southern California urban area (14). Diurnal, seasonal, and spatial variations are discussed, drawing upon a set of more than 6500 observations of ambient methylchloroform and tetrachloroethylene.
0013-936X/91/0925-1930$02.50/0
0 1991 American Chemical Society
