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Environmental Toxicology and Chemistry, Vol. 18, No. 9, pp. 1927–1931, 1999q 1999 SETAC

Printed in the USA0730-7268/99 $9.00 1 .00

EFFECTS OF A NONIONIC SURFACTANT ON BIODEGRADATION OFPHENANTHRENE AND HEXADECANE IN SOIL

RICHARD E. MACUR and WILLIAM P. INSKEEP*Department of Land Resources and Environmental Sciences, Montana State University, Bozeman, Montana 59717, USA

(Received 3 August 1998; Accepted 22 December 1998)

Abstract—The influence of a nonionic (alcohol ethoxylate) surfactant (Witconol SN70) on biodegradation of phenanthrene andhexadecane (nonaqueous-phase liquid) in soil was studied in batch and transport systems. Simultaneous enhancement of phenanthreneand hexadecane degradation was noted at surfactant doses resulting in aqueous-phase surfactant concentrations below the criticalmicelle concentration (CMC). Conversely, degradation rates of both compounds declined to essentially zero at supra-CMC doses,suggesting that distinct mechanisms of inhibition and enhancement were operating depending on the effective surfactant concentration(i.e., accounting for surfactant sorption, log KD 5 2.2 L/kg). Surfactant doses resulting in enhanced degradation correlated withenhanced gross microbial activity as determined using total CO2 evolution rates. Supra-CMC doses that resulted in inhibiteddegradation did not suppress gross microbial activity. Furthermore, measurements of phenanthrene solubilization and surface tensionindicated that phenanthrene was solubilized at supra-CMC levels of surfactant. Mechanisms of inhibition of phenanthrene andhexadecane degradation at supra-CMC surfactant concentrations may include changes in interfacial chemistry and subsequent masstransfer processes due to sorbed surfactant, reduced bioavailability of micelle-bound phenanthrene and hexadecane, or inhibitionof specific members of the microbial community responsible for hydrophobic organic compound degradation.

Keywords—Bioremediation Bioavailability Polycyclic aromatic hydrocarbons Nonaqueous-phase liquids Hydro-phobic organic compounds

INTRODUCTION

Partitioning of organic contaminants into nonpolar envi-ronments, such as natural organic matter (NOM) and non-aqueous-phase liquids (NAPLs), and subsequent mass transferlimitations during desorption often result in reduced bioavail-ability to degrading microorganisms [1]. The persistence andslow release of hydrophobic organic compounds (HOCs) fromNOM and NAPLs may represent long-term sources of pollu-tion and limit the effectiveness of bioremediation. Althoughsurfactants are useful for enhancing desorption rates and ap-parent solubilities of HOCs in soils, it remains uncertainwhether surfactants are useful for enhancing microbial deg-radation.

The use of surfactants for soil washing stems from theirability to enhance the apparent solubility of HOCs and NAPLs[2–4]. The degree of solubility enhancement varies with sur-factant structure, concentration in solution, hydrophobicity ofthe HOC, and degree of contaminant aging with sorbing ma-terial. Because surfactants can increase the apparent solubilityand transport of HOCs in soils, it is reasonable to expect thatHOC bioavailability can be increased as well. However, effortsto use surfactants for enhancing microbial degradation ofHOCs have produced varied results [5]. Some studies havereported enhanced degradation rates of organic contaminantsby nonionic surfactants at concentrations below [6,7] andabove the critical micelle concentration (CMC) [8–11], where-as other studies have shown significant inhibition of contam-inant degradation at surfactant concentrations above the CMC[12–14]. Proposed mechanisms for inhibition of microbial deg-radation at supra-CMC levels include surfactant toxicity, sur-

* To whom correspondence may be addressed(binskeep@montana.edu).

factant-enhanced contaminant toxicity, preferential use of thesurfactant as a substrate, interference of cell membrane pro-cesses, changes in cell adhesion or attachment at surface–H2O–NAPL interfaces, and reduced bioavailability of micelle-boundHOC [5,11–14]. Because of the complexity of factors con-trolling HOC bioavailability in surfactant–soil–NAPL systems(e.g., mass transfer of HOCs, cell hydrophobicity, and cellattachment at interfaces), the status of surfactant-enhanced bio-degradation in soil systems is somewhat clouded. The variedresults on surfactant-enhanced biodegradation reflect the dif-ficulty in understanding unifying themes useful for guidingbioremediation strategies [5].

Abdul et al. [15] assessed 10 commercial surfactants fortheir suitability to remove petroleum hydrocarbons from po-rous media, and Witconol SN70 (nonionic alcohol ethoxylate;Witco, Houston, TX, USA) exhibited the most favorable char-acteristics for in situ soil and aquifer washing. Furthermore,in biodegradation studies by Thibault et al. [16], WitconolSN70 added in combination with a consortium of known py-rene degraders was the most effective treatment for enhancingpyrene mineralization. The positive results reported by Thi-bault et al. [16] and Abdul et al. [15] support attempts to usenonionic surfactants such as Witconol SN70 in bioremediationefforts. The primary goal of this study was to simultaneouslyevaluate surfactant sorption, micellization, phenanthrene sol-ubility, and phenanthrene degradation as a function of surfac-tant dose in both batch and transport environments. To ac-complish this goal, the degradation of phenanthrene (repre-sentative HOC) was measured in unsaturated soil–surfactant(Witconol SN70) systems under both batch and column con-ditions in the presence of hexadecane (a representative NAPL).Our results show significant enhancement of phenanthrene andhexadecane degradation at sub-CMC surfactant doses and

1928 Environ. Toxicol. Chem. 18, 1999 R.E. Macur and W.P. Inskeep

complete inhibition of degradation at surfactant doses resultingin supra-CMC conditions.

MATERIALS AND METHODS

Soil

The surface horizon of a Wheeling loam was collected inTuscarawas County, Ohio, USA. The soil was air-dried andpassed through a 2-mm sieve before storage at 48C. The soilcontained 1.46% organic C [17], 0.2% total Kjeldahl N, 36.2mg/kg extractable P [18], 52% sand, 32% silt, and 16% clay.The 1:1 pH and electrical conductivity were 8.3 and 0.85 dS/m, respectively. Gravimetric water content at field capacity(33 kPa) was determined to be 20.9% [19]. Enumeration ofthe total culturable bacterial population yielded 8.5 3 106

CFU/g using a plate count method on nutrient rich agar [20].

Chemicals

Phenanthrene (water solubility 5 1.29 mg/L [21]), a po-lyaromatic hydrocarbon common to many petroleum-contam-inated sites, was used as a model substrate in this study.n-Hexadecane (water solubility 5 3.6 mg/L) was chosen asa representative NAPL. Radiochemicals ([1-14C]hexadecaneand [9-14C]phenanthrene) were obtained from Sigma Chem-ical (St. Louis, MO, USA). Witconol SN70, a commerciallyavailable nonionic surfactant was obtained from Witco. Wit-conol SN70 is a nontoxic [22] linear alcohol ethoxylate(CH3(CH2)x(OCH2CH2)yOH), where x 5 10 to 14 (primarily10) and y 5 5 (average molecular weight 5 392 g/mol andspecific gravity 5 0.98 g/cm3). Published values of the CMCrange from 25 to 100 mg/L [15,16]. Our own measurementsusing a conventional plot of surface tension versus log sur-factant concentration suggest a CMC of approx. 75 mg/L.

Effects of surfactant on phenanthrene solubilization

Relationships among surfactant dose, aqueous-phase sur-face tension, and phenanthrene solubilization were investi-gated in batch vessels. One-gram subsamples of Wheeling soilwere added to 25-ml Corex centrifuge tubes, followed by 20-ml additions of a mixture of nonlabeled and [14C]phenanthrenedissolved in hexadecane. The final concentration of phenan-threne in the soil was 0.1 mg/g and included 8.3 3 105 Bq/kg of [14C] phenanthrene. The final n-hexadecane concentra-tion was 15 mg/g. Surfactant doses ranging from 0 to 200mg/g were achieved by adding surfactant solutions containingdifferent concentrations of Witconol SN70 to bring the totalsolution volume to 10 ml (solid:solution ratio 5 1:10). Thetubes were shaken at 200 oscillations/min for 24 h at 22 628C and then centrifuged at 6,000 g for 20 min. A 24-h equil-ibration time was chosen because previous experimentsshowed that (1) no detectable degradation had occurred by thistime and (2) more than 96% of the added phenanthrene wassorbed within 24 h (data not shown). One milliliter of theaqueous phase was removed for analysis of [14C]phenanthrene(Tricarb CA2200 scintillation counter, Packard, Meriden, CT,USA). The remainder of the aqueous phase was removed andused for determination of surface tension with a Cenco-duNouy Interfacial Tensiometer 70545 (Central Scientific, Frank-lin Park, IL, USA).

Phenanthrene mineralization

Degradation experiments were conducted under unsaturatedsoil conditions in 30-ml glass bottles. Phenanthrene was dis-solved in hexadecane and added to 5 g of air-dried Wheeling

soil to give final concentrations of 0.1 mg/g phenanthrene and15 mg/g hexadecane containing 1.7 3 105 Bq/kg of[14C]phenanthrene or [14C]hexadecane. The solution was ap-plied dropwise from a 20-ml syringe onto soil simultaneouslymixing on a vortex shaker. Surfactant solutions ranging from0 to 200 g/L were added to the soils to bring the gravimetricwater content (um) to 0.20, yielding final surfactant concen-trations ranging from 0 to 40 mg/g soil. Humidified CO2-freeair was pumped through each vessel at a rate of approx. 2 ml/min and passed through 0.5 M NaOH to trap CO2 and 14CO2.One-milliliter aliquots of the trap solutions were analyzed for14CO2 using scintillation analysis; the remaining trap solutionwas used to determine carbonate alkalinity [23] to estimatetotal respiration rates. Rates and times of maximum phenan-threne degradation were obtained by fitting 14CO2 evolutiondata to a three-half-order kinetic equation and solving the sec-ond derivative form of the equation [24,25]. Rates of hexad-ecane degradation were determined for periods correspondingto maximum phenanthrene degradation, which usually oc-curred within 3 to 5 d. After termination of the experiments,1-g soil subsamples were combusted in a biological oxidizer(Model OX-300, R. J. Harvey Instrument, Hillsdale, NJ, USA)for analysis of residual 14C. Mass balance calculations basedon 14CO2 evolved plus residual 14C indicated phenanthrene andhexadecane recoveries of 93 6 10% and 97 6 9%, respec-tively.

Phenanthrene mineralization under transport conditions

Phenanthrene degradation was also studied under unsatu-rated transport conditions using polycarbonate columns (length5 55 mm and width 5 32 mm) packed with 20% Wheelingloam uniformly mixed with 80% acid-purified quartz sand (40–100 mesh, Fluka Chemical, Milwaukee, WI, USA) to a bulkdensity of 1.30 6 0.05 g/cm3. A constant potential of 30 kPawas applied to a porous fiberglass membrane (bubbling pres-sure ø 100 kPa) located in the bottom end cap. Influent wasapplied to the top of the columns with a continuous flow pumpset to deliver 0.91 ml/h (approx. 1 pore volume/d; flux density5 0.11 cm/h and pore-water velocity 5 0.33 cm/h). Columnswere conditioned by eluting 10 pore volumes of 5 mM KCland then treated with 1.2 mg of phenanthrene (with 7.4 3 104

Bq of 14C-phenanthrene) dissolved in 185 mg of hexadecane(applied dropwise to the soil surface). The amounts of phen-anthrene and hexadecane applied per unit mass of soil wereidentical to concentrations used in batch degradation experi-ments. Air was delivered (0.8 ml/min) to the lower end of thecolumn and passed through 0.5 M NaOH at the top exit as amechanism to minimize O2 depletion and to trap 14CO2 releasedfrom phenanthrene degradation. CO2-trap solutions (0.5 MNaOH) were also placed in the vacuum chamber, exchangedat 2-d intervals, and analyzed for 14C as described above.

Total 14C in column effluent samples was determined byremoving 1.0-ml aliquots for scintillation analysis. An addi-tional 1.0-ml aliquot was acidified and sparged with N2(g) torelease dissolved [14C]carbonate and reanalyzed for 14C. Theremaining effluent was used to measure surface tension asdescribed above. Selected samples were analyzed usingHPLC–radioisotope detection (mobile phase 5 80% CH3CN,20% H2O, and 0.1% H3PO4; flow rate 5 1.0 ml/min; 250- 34.6-mm C18 10-mm column) to determine the distribution of14C species in the aqueous phase. No 14C-labeled peaks otherthan 14C-phenanthrene were observed (detection limit ø 30Bq/ml) in the effluent. After termination, the columns were

Surfactant effects on biodegradation of phenanthrene in soil Environ. Toxicol. Chem. 18, 1999 1929

Fig. 1. Maximum phenanthrene (PHEN) degradation rates (Rmax, m),hexadecane (HEX) degradation rates (v), and CO2 evolution rates(m) as a function of surfactant dose. All rates are calculated valuesoccurring at the time of maximum phenanthrene degradation. CMC95 surfactant dose necessary to achieve the aqueous-phase criticalmicelle concentration. Symbols to the left of dashed lines indicatevalues at a surfactant dose of 0 mg/g. Error bars are SEMs.

Fig. 2. Phenanthrene (PHEN) solubilization (v), aqueous-phase sur-face tension (m), and corresponding phenanthrene degradation rates(Rmax, m) as a function of surfactant dose. CMC9 5 surfactant dosenecessary to achieve the aqueous-phase critical micelle concentration(;75 mg/L). Symbols to the left of dashed lines indicate values at asurfactant dose of 0 mg/g. Error bars are SEMS.

frozen, sectioned, and analyzed for total residual 14C by com-busting 1.0-g samples in a biological oxidizer, followed byliquid scintillation. Mass balance calculations based on 14Celuted, 14CO2 evolved, and residual 14C yielded a 14C recoveryof 67.2% for the column experiment presented here.

RESULTS AND DISCUSSION

Degradation under batch conditions

Additions of Witconol SN70 at low concentrations (0.2–2.0 mg/g) significantly enhanced (p 5 0.05) the rates of phen-anthrene and hexadecane degradation relative to control treat-ments containing no surfactant (Fig. 1). Degradation rates ofboth compounds were reduced at surfactant doses .5 mg/gand completely inhibited at doses .10 mg/g. To understandthe concentration-dependent effects of Witconol SN70 onphenanthrene and hexadecane degradation, we also determinedthe concomitant effects of surfactant dose (S0) on gross mi-crobial activity. Total microbial respiration rates in the absenceof surfactant averaged 500 nmol/h per g (Fig. 1). A pronouncedpeak in CO2 evolution rates (1,000 nmol/h per g) occurred atS0 5 2 mg/g, corresponding to points of maximum phenan-threne and hexadecane degradation rates. Carbon budget cal-culations reveal that increased CO2 evolution rates at S0 5 2mg/g were due primarily to increases in hexadecane degra-dation rates. At S0 near 5 mg/g, the rates of phenanthrene andhexadecane degradation declined significantly (p 5 0.05);however, CO2 evolution rates returned to values near thoseobserved in the absence of surfactant. Total CO2 evolutionrates then increased dramatically at S0 . 20 mg/g while HOCdegradation was still inhibited, suggesting mineralization ofsurfactant by indigenous microorganisms. These results areconsistent with those of other studies showing that WitconolSN70, and more generally alcohol ethoxylates, are degradableunder aerobic conditions [16,22]. Furthermore, data on CO2

evolution rates demonstrate that higher surfactant doses (orthe presence of compounds solubilized by surfactant appli-cations) did not inhibit gross microbial respiration. However,our observations do not preclude the possibility that surfactantdoses .2 mg/g adversely affected specific phenanthrene or

hexadecane degraders, which represented only a fraction ofthe total microbial population.

Relative to simple aqueous systems, effects of surfactantson microbial degradation of organic contaminants in soil-watersystems are complicated because of surfactant sorption priorto micellization in the aqueous phase. To help visualize therelationship among surfactant dose (S0), micellization, phen-anthrene solubilization (sorption) and maximum phenanthrenedegradation rates (Rmax, [24,25]), results from the batch deg-radation studies and batch solubilization experiments wereplotted together as a function of surfactant dose (Fig. 2). Thesurface tension data exhibit three distinct, nearly linear seg-ments whose characteristics are governed by sorption and mi-cellization processes. Nonionic surfactant sorption onto soil atlow solution concentrations is Freundlich-type [26], duringwhich surfactant additions result in relatively small increasesin solution concentration (as indicated by minimal changes insurface tension). Further increases in surfactant dose (S0 50.2–10 mg/g) resulted in an intermediate region of log-linearsorption (also Freundlich-type) where surface tension decreas-es. At S0 . 10 mg/g, surface tension reaches a minimum valuecorresponding to the formation of surfactant micelles. Thesurfactant dose necessary to achieve the aqueous-phase CMCis referred to here as CMC9 and accounts for surfactant sorptionby soil before micellization. The CMC9 determined in ourexperiments was approx. 13 mg/g. The amount of surfactantsorbed can be estimated by assuming that the aqueous-phasesurfactant concentration (SAQ) at CMC in the presence of soilwas similar to SAQ at the CMC in soil-free water (75 mg/L).At CMC9 5 13 mg/g, approx. 90% of added surfactant issorbed, yielding an estimated sorption coefficient (log KD) of2.2 6 0.2 L/kg. This value is within a factor of 4 of KD valuespreviously reported for an alcohol ethoxylate with a structuresimilar to that of Witconol SN70 (log KD 5 1.60–1.79 L/kg[26]).

Data on the amount of phenanthrene solubilized as a func-tion of surfactant dose were consistent with surface tensiondata and showed significant increases in apparent phenanthrenesolubility above the effective CMC of 13 mg/g. Increases inphenanthrene and hexadecane degradation rates occurred at

1930 Environ. Toxicol. Chem. 18, 1999 R.E. Macur and W.P. Inskeep

Fig. 3. Percentage of initial phenanthrene degraded (m) or eluted (m)and effluent surface tension (v) as a function of time in an unsaturatedtransport environment. Influent surfactant concentrations were in-creased stepwise from 0 (region I) to 200 (region II), 2,000 (regionIII), and 10,000 mg/L (region IV).

surfactant concentrations well below that necessary to achievethe CMC (i.e., at S0 5 0.2–2 mg/g; Fig. 2). Consequently, anyenhancement in the degradation rate of phenanthrene or hexa-decane was not due to solubility enhancement as determinedin the bulk aqueous phase. Conversely, the inhibition of phen-anthrene and hexadecane degradation rates correlated well withthe onset of micellization and the corresponding increase inapparent phenanthrene solubility.

Our results are consistent with recent reports that micelli-zation may have profound effects on phenanthrene bioavail-ability to degrading microorganisms. In studies using a mixedenrichment culture, Guha and Jaffe [14,27] showed that in-hibition of phenanthrene degradation in aqueous–surfactantsystems (no soil) at CMC and supra-CMC doses was associatedwith reduced bioavailability due to partitioning of phenan-threne into surfactant micelles. Model estimates of the bio-available fraction of micelle-bound phenanthrene (f in Guhaand Jaffe [14,27]) varied with surfactant structure. For Brij 35(C12H25(OCH2CH2)23OH), f was essentially 0 across a widerange of surfactant concentrations; for Brij 30(C12H25(OCH2CH2)4OH), f ranged from near 1 to near 0 withincreasing doses above the CMC. In our experiments, phen-anthrene degradation rates dropped to essentially 0 after mi-cellization despite the fact that phenanthrene was solubilizedby the surfactant (i.e., no longer sorbed). Although these datamight suggest that the bioavailable fraction of micelle-boundphenanthrene was essentially zero in our experiments, it isimportant to note that in soil–water systems, a significantamount of surfactant is sorbed by soil particles before micel-lization. In the current experiments, more than 90% of addedsurfactant was sorbed at the effective CMC of 13 mg/g. It isunknown whether sorbed surfactant may influence cell attach-ment (adhesion) or contaminant mass transfer rates and sub-sequently play a role in the inhibition of HOC degradation.

Phenanthrene degradation under transport conditions

Column studies were conducted to determine whether re-lationships among surfactant dose, micellization, phenanthrenesolubilization, and microbial degradation rates were consistentbetween batch and transport environments. Influent surfactantconcentrations were increased in a stepwise manner during theexperiment, resulting in three distinct surface tension plateaus(Fig. 3). Plateaus in the surface tension curves suggest thataqueous-phase surfactant concentrations attained equilibriumvalues dependent on influent surfactant concentrations (S0). Inregion I, before surfactant addition, the majority of phenan-threne was sorbed (essentially no phenanthrene elution), and,after a brief lag phase, phenanthrene degradation was nearlylinear, corresponding to a rate of approx. 0.19 nmol/h per g.In region II (200 mg/L), surfactant sorption reduced inputsurfactant concentrations to below the CMC, as indicated byonly a partial drop in surface tension from 70 to 50 dynes/cm(surface tension at CMC ø 25 dynes/cm). Very little phen-anthrene was solubilized and eluted during this period becausethe aqueous-phase surfactant concentrations were below theCMC. However, phenanthrene degradation rates increasedfrom 0.19 to 0.95 nmol/h per g. These observations are similarto, but more pronounced than, results from batch experiments,in which sub-CMC surfactant levels increased phenanthreneand hexadecane degradation. In region III (2,000 mg/L), sur-face tension decreased to values corresponding to surfactantconcentrations greater than the CMC. Phenanthrene solubili-zation becomes apparent in region III (1,000 h); however,

phenanthrene degradation rates declined. At this stage of theexperiment, it can be argued that the phenanthrene potentiallyavailable for degradation was nearly exhausted. An additionaltransport experiment in which 2,000 mg/L surfactant was con-tinuously added from the start of the experiment showed nearlycomplete inhibition of phenanthrene degradation at supra-CMC levels of surfactant (SAQ . CMC, data not shown). In-hibited phenanthrene degradation at surfactant concentrationsgreater than the CMC were consistent with data shown inFigure 2 for batch environments. Finally, in region IV (10,000mg/L), a distinct increase in phenanthrene solubilization andelution was noted at 1,300 h (Fig. 3), corresponding to furtherincreases in micelle concentration.

In summary, degradation rates of phenanthrene and hexa-decane were enhanced at sub-CMC levels but significantlyinhibited above total surfactant concentrations necessary toachieve CMC. These results indicate that information on theeffective surfactant concentration relative to CMC (accountingfor sorption) is critical to designing and interpreting resultsfrom experiments with surfactants in sorptive environments.Furthermore, the results suggest that two discrete mechanismswere functioning at the different surfactant concentrations.Low surfactant doses may have enhanced degradation andgross microbial activity by improving cell–soil–NAPL contact,resulting in increased mass transfer rates and/or increasedmembrane fluidity with subsequent increases in contaminantuptake rates [5–7]. Our results also showed that inhibition ofdegradation above CMC was not due to gross toxicity of Wit-conol SN-70 or solubilized HOC to soil microorganisms. Twodistinct changes in physiochemical conditions occur upon mi-cellization in soil systems that may play a role in the inhibitionof HOC degradation. First, partitioning of phenanthrene intosurfactant micelles may result in reduced phenanthrene bio-availability, consistent with observations made in aqueous–surfactant systems [14,27]. Second, the amount of surfactantsorbed before the CMC is achieved is very significant in soilsystems. Surfactant saturation of sorbing phases within soilmay influence surface hydrophobicity, which could play an

Surfactant effects on biodegradation of phenanthrene in soil Environ. Toxicol. Chem. 18, 1999 1931

important role in cell adhesion or attachment or mass transferof HOCs across interfaces.

Acknowledgement—This work was supported in part by the GreatPlains Hazardous Substances Research Center, project 94-09; theArmy Corps of Engineers Waterways Experiment Station, projectW81EWF4318-0038; and the Montana Agricultural Experiment Sta-tion, project 104398.

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