review of natural attenuation of btex and mtbe in groundwater

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Review of Natural Attenuation of BTEX and MTBEin Groundwater

Eric A. Seagren, A.M.ASCE,1 and Jennifer G. Becker2

Abstract: Monitored natural attenuation has become accepted as an appropriate ‘‘technology’’ for treating petroleum hydrocontaminated groundwater. Commonly, the monoaromatic compounds benzene, toluene, ethylbenzene, and xylenes~BTEX! and oxygen-ated additives, e.g., methyltert-butyl ether~MTBE!, are the major constituents of regulatory importance. Although a variety of natuoccurring attenuation mechanisms may reduce BTEX and MTBE concentrations, intrinsic biodegradation is the primary demechanism. Attenuation by anaerobic biodegradation is important for BTEX compounds because of the relative abundance ofelectron acceptors as compared with dissolved oxygen. Field studies indicate that MTBE is also biodegradable in shallow aquifea slower rate relative to BTEX compounds; thus, dispersion and dilution may also be important MTBE attenuation mechandemonstrate that natural attenuation is occurring, it is critical to document the proposed natural attenuation processes in the fiprocesses often cause measurable ‘‘footprints.’’ At a large number of petroleum-hydrocarbon release sites, natural attenuationhave been observed and documented and found to control the extent of migration of BTEX contaminants. Although the presencemay compromise the use of natural attenuation at some sites due to the greater mobility and persistence of MTBE comparedsome data indicate that MTBE plumes do eventually stabilize.

DOI: 10.1061/~ASCE!1090-025X~2002!6:3~156!

CE Database keywords: Attenuation; Biodegradation; Case reports; Groundwater.

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Introduction

During the 1980s, significant effort was expended to assessmonitor the contamination of groundwater by petroleum hydcarbon releases in the United States~Chapelle 1999!. Numerouscase studies at field sites indicated that contaminant plumesulting from releases of petroleum hydrocarbons in groundweither migrated more slowly than expected, reached a steady sor decreased in extent and concentration over time. Plume mtion was inhibited due to naturally occurring attenuation mecnisms, especially biodegradation, but also dilution, dispersand sorption, e.g.~Barker et al. 1987; Kemblowski et al. 1987Chiang et al. 1989; Barbaro et al. 1992; Baedecker et al. 19McAllister and Chiang 1994; Wiedemeier et al. 1996!. Use ofthese natural attenuation processes as a remedial approachmanaged strategy of monitored natural attenuation~or intrinsicremediation! has become accepted as an appropriate ‘‘techogy’’ for treating petroleum hydrocarbon contaminated grounwater.

1Assistant Professor, Dept. of Civil and Environmental Engineeri1149 Martin Hall, Univ. of Maryland, College Park, MD 20742~corre-sponding author!. E-mail: [email protected]

2Assistant Professor, Dept. of Biological Resources Engineering, 1Agricultural Engineering Building, Univ. of Maryland, College ParMD 20742. E-mail: [email protected]

Note. Discussion open until December 1, 2002. Separate discusmust be submitted for individual papers. To extend the closing dateone month, a written request must be filed with the ASCE ManagEditor. The manuscript for this paper was submitted for review and psible publication on March 1, 2002; approved on March 1, 2002. Tpaper is part of thePractice Periodical of Hazardous, Toxic, and Radioactive Waste Management, Vol. 6, No. 3, July 1, 2002. ©ASCE, ISSN1090-025X/2002/3-156–172/$8.001$.50 per page.

156 / PRACTICE PERIODICAL OF HAZARDOUS, TOXIC, AND RADIOACT

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The goal of this article is to review the state of the artremediation of petroleum hydrocarbon releases using naturatenuation. Although many hydrocarbons are present in gasoand other refined petroleum products, this article focuses on nral attenuation of the monoaromatic compounds benzene, toluethylbenzene, and xylenes~BTEX! and the gasoline additive methyl tert-butyl ether ~MTBE!, because they are commonly thmajor constituents of regulatory importance at petroleum relesites~McAllister and Chiang 1994; ASTM 1998!. Specifically, wereview the key attenuation mechanisms in groundwater, dischow natural attenuation of BTEX and MTBE is demonstrateand summarize the findings from the field.

Attenuation Mechanisms for BTEX and MTBE

Once petroleum hydrocarbons are released into the environmthe nonaqueous phase liquid~NAPL! source material may volatilize into the soil gas and/or dissolve into the aqueous phaConstituents such as BTEX compounds and MTBE that havesolved into the aqueous phase may subsequently be transpwith the groundwater and are also subject to several possnatural attenuation mechanisms, including physical phenomchemical reactions, and biological processes.

Physical Mechanisms

There are several physical processes that may contribute tonatural attenuation of petroleum hydrocarbons. These are alldestructive attenuation mechanisms that may result in a decrin contaminant concentration, but do not reduce the total contanant mass. The key physical natural attenuation processes incdispersion, sorption, and volatilization, which are discusseddetail below. Infiltration and recharge are not reviewed here,

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though these processes may also impact natural attenuatiodilution and transfer of materials~e.g., electron donors and accetors or contaminants! from the unsaturated zone to the saturazone.

DispersionDispersion and diffusion are the only mixing processes for soluin the deep subsurface and, thus, can play a key role in mixingBTEX and MTBE contaminant plume with electron acceptopresent in the surrounding aquifer~Cho et al. 1997!. In addition,dispersion and diffusion can play a key role in supplying electacceptors to the contaminated zone. For example, oxygen ingroundwater upgradient of the contaminated zone is transpoby advection into the contaminated zone~Fig. 1!. Because theconsumption of oxygen is relatively fast, and the supply of ogen is slow due to the low oxygen solubility in water, microoganisms near the source area may consume the available oxto biodegrade a portion of the hydrocarbons. In this case,remaining hydrocarbons are transported downgradient inoxygen-limited plume~Borden 1994!. If this occurs, oxygentransport via dispersion from the surrounding groundwatercomes an important mode of oxygen supply for hydrocarbon dradation downstream of the source~Kemblowski et al. 1987; Bor-den et al. 1997!, resulting in aerobic biodegradation occurrinprimarily at the plume fringes~Borden 1994!. In addition, supplyof oxygen via transfer across the capillary fringe/water tabletransverse dispersion may be important for aerobic biodegrtion of light NAPLs~LNAPLs! at or near the water table. Becaudispersion is a relatively limited process, the overall aerobic bdegradation rate may be slow, and oxygen availability may pan important role in determining reaction rates and hydrocarpersistence~Barker et al. 1987!.

Hydrodynamic dispersion may also result in dilution of tcontaminant concentrations. This may be a particularly imporattenuation mechanism for MTBE, because of its apparent recitrance to biodegradation under some conditions and its tendto remain in the aqueous phase and move unretarded withgroundwater flow~Anthony et al. 1999!. For example, based othe occurrence of recharge events and a low biodegradationtential, Landmeyer et al.~1998! concluded that the observed dcreases in MTBE concentrations relative to benzene in a shagasoline-contaminated aquifer were the result of dilution andpersion. However, a modeling analysis at the Borden field sOntario, using measured dispersivity values and pseudo-fiorder biodegradation rates based on the field mass loss of Mand BTEX compounds suggested that dispersion alone is nvery effective natural attenuation process~Schirmer et al. 1999!.

Fig. 1. Oxygen supply to petroleum-hydrocarbon release undergonatural attenuation@adapted from Figs. 9.1 and 9.2, p. 184, in Bord~1994!#

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Relatively high, near-source concentrations of MTBE also incate that dispersion may be insufficient for receptor protectionmany cases~Schirmer and Barker 1998!.

SorptionSorption may impact natural attenuation by affecting a contanant’s migration relative to the groundwater flow and its dissolvconcentration, which in turn affects the rates of other transpand transformation reactions, e.g., volatilization and biodegration ~ASTM 1998!. Based on batch sorption isotherms, the BTEcompounds have different sorption tendencies~Kemblowski et al.1987; Stuart et al. 1991; Zytner 1994!. As a result, the individualBTEX compounds have been observed to migrate at differvelocities in laboratory column studies~Stuart et al. 1991; Angleyet al. 1992; Larsen et al. 1992!. This same ‘‘chromatographicseparation’’ of BTEX occurs in petroleum-hydrocarbon impactgroundwater aquifers~Odermatt 1994!, thereby affecting the relative spatial and temporal proportions of these compounds inute plumes.

Retardation factors~R! for BTEX based on published organicarbon partition coefficients (Koc), linear and reversible sorptionand assumed aquifer properties are given in Table 1.predicted differences in mobility @benzene.toluene.(ethylbenzene and xylenes)# could explain the observed chromatographic separation of the BTEX compounds. This predicorder of relativeR values is consistent withR values estimated byPtacek et al.~1987! using different methodologies for the BTEXcompounds in a sandy aquifer of low organic carbon contebenzene, 1.0–1.4; toluene, 1.2–1.5; ando-, m- andp-xylene, 1.3–2.0. Because the aquifer material had a relatively low orgacontent (f oc'0.02%), theseR values are significantly smallethan the values in Table 1. These low values indicate thatgeneral, sorption of the BTEX compounds in groundwater mayassumed to be relatively limited in aquifer materials with loorganic carbon content~Chiang et al. 1989!. Of course, BTEXsorption could be more significant in higher organic content sosuch as peat moss or organic top soil~Zytner 1994!, and sorptionof BTEX onto inorganic surfaces can also occur~Rogers et al.1980!.

Using the same assumptions as above for the BTEX copounds, the estimatedR value for MTBE~Table 1! is 1.70, whichis consistent withR values near 1 that were estimated in fiestudies~Borden et al. 1997; Schirmer and Barker 1998!. Thus,sorption is generally an insignificant natural attenuation procfor MTBE. As a result, MTBE migrates more rapidly than thBTEX compounds, causing a chromatographic-like separatioMTBE and the BTEX compounds in contaminant plumes~Land-meyer et al. 1998!. The greater susceptibility of BTEX compounds to biodegradation as compared with MTBE~discussedbelow! further amplifies this separation effect. Therefore, tleading edge of a plume emanating from a spill of MTBoxygenated gasoline may contain substantial levels of MTwith little or no BTEX ~Squillace et al. 1997!.

VolatilizationVapor pressures and Henry’s Law constants, which govern elibrium partitioning from the NAPL and groundwater, respetively, are provided in Table 1 for the BTEX compounds aMTBE. Although partitioning can occur from the petroleum prouct NAPL and groundwater, the focus here is on volatilizatifrom the dissolved contaminant plume.

A larger Henry’s constant indicates a greater tendency forchemical to volatilize. The Henry’s constants for all of the BTE

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Table 1. Physical/Chemical Properties of BTEX and MTBE~Value at 25°C!

Compound

Watersolubility~mg/L! Log Kow Log Koc

Vaporpressure~mm Hg!

Henry’sconstant,KH ,~atm m3/mol!

Henry’sconstant,

KH8 ~dimensionless!Retardationfactor,Ra

Cgasb

~mg/L!

Benzene 1,791c 2.13c 1.09–2.53d 95.19c 5.43(1023)c, 5.28(1023)e 2.22(1021)f, 2.16(1021)e 1.76–22.0 13.4Toluene 534.8c 2.73c 1.12–2.85d 28.4c 5.94(1023)c, 6.43(1023)e 2.43(1021)f, 2.63(1021)e 1.82–45.9 7.4Ethylbenzene 161g 3.15g 1.98–3.04d 9.53g 8.44(1023)g, 7.78(1023)e 3.45(1021)f, 3.18(1021)e 6.92–69.0 1.3

o-Xylene 175c 3.12c 1.68–2.73d 6.6c 5.1(1023)c, 4.99(1023)e 2.084(1021)f, 2.04(1021)e 2.97–34.3 1.2

m-Xylene 146c 3.20c 2.04–3.15d 8.3c 7.68(1023)c 3.139(1021)f 7.80–88.6 3.0

p-Xylene 156c 3.15c 2.05–3.08d 8.7c 7.68(1023)c 3.139(1021)f 7.96–75.5 1.2

MTBE 48,000–54,300h,i 1.24h 1.05f 249h 5.87(1024)h, 5.28(1024)e 2.399(1022)f, 2.16(1022)e 1.70 87.6aCalculated based on linear sorption,R511(rb /n)KD , KD5 f ocKoc , Koc values in column 4, and assumingrb51.86 g•cm23, n50.3, and f oc

50.01.bEquilibrium gas-phase concentration, calculated using Henry’s Law,Cg5KH8 C, KH8 values in column 7, and assuming equilibration with the followiaverage aqueous concentrations in equilibrium with PS-6 gasoline with 10%~volume! MTBE: benzene, 60.5 mg/L; toluene, 30.5 mg/L; ethylbenzene,mg/L; o-xylene, 5.9 mg/L;m-xylene, 9.6 mg/L;p-xylene, 3.7 mg/L; and MTBE, 3,650 mg/L~Poulsen et al. 1992!.c Sage et al.~1990!.d Mackay et al.~1992!.e Robbins et al.~1993!.f Squillace et al.~1997!. Dimensionless Henry’s constants generally were calculated from documented references.g Jarvis et al.~1989!.h Michalenko et al.~1993!.i Mackay et al.~1993!.

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compounds are fairly similar and low relative to other petrolehydrocarbons. However, the gas-phase contaminant concetions that result from equilibration with an aqueous solutionalso a function of the aqueous contaminant concentrationthus, may vary from compound to compound~Table 1!. Becauseof the relatively low Henry’s constants and mass-transfer limtions at the capillary fringe, the actual volatilization of BTEcompounds from groundwater is expected to be relatively limi~McAllister and Chiang 1994!. For example, based on a mabalance analysis, only 5% or less of the total mass loss of benat a petroleum-product impacted sandy aquifer could be attribto volatilization ~Chiang et al. 1989!. Nevertheless, volatilizationmay account for a larger fraction of the total mass loss if:~1! thewater table is shallow or fluctuates significantly~ASTM 1998!;~2! the groundwater temperature is relatively high, whichcreases the value of the Henry’s constant~Robbins et al. 1993!; or~3! rates of biodegradation are relatively slow~McAllister andChiang 1994!.

Dimensionless Henry’s constants,KH8 , for MTBE ~Table 1!are an order of magnitude lower than theKH8 of the BTEX com-pounds. TheKH8 values for other alkyl ether oxygenates are simlar in magnitude; therefore, MTBE and other fuel oxygenates ptition to a high degree into water~Squillace et al. 1997!.However, given the higher aqueous solubility of MTBE as copared to the BTEX compounds, the equilibrium gas-phase ccentration could be higher than that for the BTEX compoun~Table 1!.

Mass losses of BTEX and MTBE from groundwater via votilization may be further reduced for portions of plumes that.1 m below the water table~Schirmer and Barker 1998!. This isdue to the weakness of vertical transverse dispersion ingroundwater and the downward groundwater velocity in thecinity of the water table during recharge~Rivett 1995!.



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Chemical Mechanisms

BTEX compounds are not expected to be transformed by nrally occurring chemical reactions under typical groundwater cditions ~McAllister and Chiang 1994!. Although benzene hasmolecular formula indicating a high degree of unsaturation, bzene is highly stable and resistant to the reactions characterisother unsaturated compounds, e.g., alkenes~Morrison and Boyd1973!. The stability of benzene and other aromatic compounddue to several factors—in particular, the delocalization of thepelectrons and the manner in which the various orbitals that mup thep cloud are filled. The stability and low reactivity of benzene and its derivatives is borne out by observations of little orBTEX mass loss in aerobic and anaerobic batch~Barker et al.1987; Kemblowski et al. 1987; Chiang et al. 1989; Baedecet al. 1993!, and column experiments~Angley et al. 1992; Larsenet al. 1992!, in which biological activity was prevented.

Similarly, transformation of MTBE by naturally occurrinchemical reactions under typical groundwater conditions isexpected~Squillace et al. 1997; Schirmer and Barker 1998!. Ingeneral, ethers are relatively unreactive compounds~Morrisonand Boyd 1973!. This is due to the ether linkage, which is stabtoward bases, oxidizing agents, and reducing agents, anknown to undergo only one type of chemical reaction, i.e., cleage by acids under very vigorous conditions. For example,and Novak~1995! observed chemical oxidation of MTBE totert-butyl alcohol~TBA! and acetone in soil microcosms and in soltion when H2O2 was added in the presence of catalytic ferroiron. However, the reaction was favored at low pH~pH 4! and didnot occur after oxidation and precipitation of the iron occurreTherefore, the reaction is unlikely to occur in aerobic and neneutral to alkaline groundwater environments~Squillace et al.1997!. Indeed, MTBE has been shown to be chemically unretive under aerobic and anaerobic conditions in the sterile cont


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used in numerous batch experiments~Yeh and Novak 1994; 1995Borden et al. 1997; Bradley et al. 1999!.

Biological Mechanisms—BTEX

The primary destructive attenuation mechanism for BTEX aMTBE is biodegradation. BTEX-degrading bacteria appear toubiquitous in the environment and are estimated to be preseapproximately 99% of contaminated sites~Norris 1994!.

Occurrence of BTEX Oxidation under Different TerminalElectron Accepting Process„TEAP… ConditionsAs discussed above, significant aerobic biodegradation ofsolved BTEX compounds and LNAPLs may occur at the plufringes and at or near the water table, respectively. Cometabbiotransformation of BTEX compounds can be mediated by aebic organisms that are growing on aliphatic hydrocarbons~Ritt-mann et al. 1994!. However, BTEX compounds can also be usas growth substrates by a wide variety of bacterial strains, anis generally assumed that growth-related transformationsmuch more important than cometabolic reactions in bringabout the removal of subsurface BTEX contamination.

The flow of electrons from a BTEX compound to an electracceptor releases free energy (DGR). Some of this free energy iused by the bacteria for cell synthesis~McCarty 1972!. As DGR

values become less negative, less energy is available for cellthesis. The paradigm for BTEX biodegradation in subsurfacevironments is that contaminant removal initially occurs via aebic respiration because this coupling releases the maximumenergy. After oxygen is depleted, oxidation of the BTEX copounds is coupled to the reduction of electron acceptors that yprogressively less free energy~less negativeDGR values!. Ac-cording to this paradigm, electron acceptors are depleted byoxidation of BTEX compounds in the following order: oxygeFe~III !, nitrate, sulfate, and carbon dioxide.

However, utilization of electron acceptors for the oxidationBTEX compounds may not always proceed as predicted onbasis of thermodynamic considerations. The dominant TEAP mvary both temporally and spatially in a contaminated aqu~Vroblesky and Chapelle 1994!, and in some cases, TEAPs manot be mutually exclusive. Furthermore, at a given site, biodedation of certain BTEX compounds may not occur under soredox conditions~Cho et al. 1997!, even though the potential fobiodegradation has been demonstrated in laboratory studies.degradation potential may not be realized in the field due toabsence of the appropriate electron acceptor or organisms/geinformation, or due to inhibitory environmental conditions. Neertheless, the pool of anaerobic electron acceptors in subsuenvironments contaminated with petroleum hydrocarbons is tcally much larger than the supply of oxygen, as discussed ab~Fig. 1!. Therefore, greater amounts of dissolved BTEX copounds may be removed from contaminated groundwater pluvia anaerobic metabolic processes, as compared with aerobicpiration ~Wiedemeier et al. 1999!.

When subsurface environments become anaerobic, Fe~III ! isoften one of the most abundant electron acceptors availablemicrobial metabolism~Lovley and Lonergan 1990!. Geochemicalchanges occurring concomitantly with the oxidation of BTEcompounds suggest that dissimilatory iron reduction may contute significantly to the natural attenuation of some aquifers pluted with BTEX and other organic contaminants~Lovley et al.1989; Cozzarelli et al. 1990; Lyngkilde and Christensen 19Borden et al. 1985; Rugge et al. 1995; Lovley 1997!. Although



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benzene often appears to be recalcitrant under denitrifying cotions, oxidation of benzene under Fe~III !-reducing conditions hasfrequently been demonstrated in both laboratory and field stu~Lyngkilde and Christensen 1992; Rugge et al. 1995; Kazuet al. 1997!.

Denitrification may be an important mechanism for removisome BTEX compounds in anaerobic subsurface environmthat are also impacted by agricultural activities and, therefoexhibit relatively high groundwater nitrate concentrations. In geral, under denitrifying conditions, oxidation of toluene, ethylbezene, andm- and p-xylene is reported to occur rather rapid~Zeyer et al. 1986; Hutchins 1991; Borden et al. 1997!, whileremoval ofo-xylene occurs relatively slowly~Kuhn et al. 1985;Hutchins 1991; Borden et al. 1997!. The fate of benzene undedenitrifying conditions is less certain. Although oxidation of bezene to carbon dioxide has been linked to nitrate reductionenrichment cultures developed from soil and groundwater~Bur-land and Edwards 1999!, in most other laboratory~Kuhn 1988;Hutchins et al. 1991; Kazumi et al. 1997! and field studies~Bar-baro et al. 1992; Borden et al. 1997!, benzene appeared to bbiologically recalcitrant under denitrifying conditions.

Relatively high levels of sulfate occur naturally in somgroundwater systems, including groundwaters that are influenby the dissolution of sulfate-bearing minerals and aquifers thatproximal to marine or estuarine systems. In general, when sureduction is a dominant TEAP, toluene is one of the most rapdegraded BTEX compounds in many monitored field sites~Actonand Barker 1992; Beller et al. 1995; Borden et al. 1995; Chapet al. 1996a; Davis et al. 1999!, as well as in some laboratormicrocosms and column reactors~Edwards et al. 1992; Barlazet al. 1995!. The xylene isomers are also degraded relatively ridly in BTEX mixtures under sulfate-reducing conditions. However, ethylbenzene~Edwards et al. 1992; Barlaz et al. 199Beller et al. 1995; Borden et al. 1995; Davis et al. 1999! and, to alesser extent, benzene~Edwards et al. 1992; Thierrin et al. 1993Bellar et al. 1995! sometimes appear to be recalcitrant or onminimally biodegradable when present in BTEX mixtures undsulfate-reducing conditions. The biodegradation of benzenealso often preceded by extended lag periods when it is presethe sole organic growth substrate in sulfate-reducing systems~Ka-zumi et al. 1997; Gieg et al. 1999!, perhaps because benzendegrading organisms are absent or present only in low numbethe sulfate-reducing zones of some aquifers~Weiner and Lovley1998a!.

It should be noted that, in some field and mixed-culture laratory studies, biogeochemical evidence suggests that in addto sulfate reduction, another TEAP, such as Fe~III !-reduction~Borden et al. 1995; Gieg et al. 1999! or methanogenesis~Actonand Barker 1992; Gieg et al. 1999! occurred to a significant extent and may have contributed to the oxidation of BTEX copounds in some systems. Nevertheless, pure cultures of sulreducing strains have been isolated using toluene oro- orm-xylene as organic substrates~Heider et al. 1998!, which dem-onstrates that the potential for oxidation of some of the BTEcompounds coupled to the reduction of sulfate does exist. Sevstudies conducted using enrichment cultures also provide stevidence that benzene~Lovley et al. 1995; Phelps et al. 1996! andtoluene~Beller et al. 1992; Edwards et al. 1992! mineralization isdependent on sulfate reduction.

Evidence collected in several field studies suggests that atsome of the BTEX compounds may be transformedin situ undermethanogenic conditions~Reinhard et al. 1984; Wilson et a1990; Gieg et al. 1999!. Consistent with observations of BTEX


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Table 2. Stoichiometric Requirement of Benzene Oxidation Plus Cell Synthesis for Various Electron Acceptorsa

Terminal electron acceptor Reaction Mass ratio of electron acceptor to benezene~g/g!

O2 C6H613.07 O210.89 HCO3210.89 NH4

1

→0.89 C5H7O2N12.45 CO212.12 H2O1.26:1~2.08:1!b

Fe31 C6H6112.54 Fe3110.87 HCO3210.87 NH4

114.14 H2O→0.87 C5H7O2N112.54 Fe2112.52 CO2112.54 H1

8.97:1

NO32 C6H612.54 NO3

210.86 HCO3210.86 NH4

112.55 H1

→0.86 C5H7O2N11.27 N212.55 CO213.39 H2O2.02:1

SO422 C6H613.42 SO4

2210.12 HCO3210.12 NH4

115.16 H1

→0.12 C5H7O2N11.71 H2S11.71 HS215.49 CO2

12.88 H2O

4.21:1

CO2 C6H610.09 HCO3210.09 NH4

114.20 H2O→0.09 C5H7O2N13.57 CH412.13 CO2

Not applicable

a McCarty’s ~1972! assumptions regarding efficiency of energy capture and transfer are implicit, and electron acceptor requirements of cell mawere neglected, in these determinations.bValue in parentheses includes coreactant requirement for O2 under aerobic conditions.

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biodegradation under other anaerobic electron-accepting cotions, toluene generally appears to be the most readily degradBTEX compound under methanogenic conditions~Wilson et al.1986!. In some cases, recalcitrance of benzene~Acton and Barker1992; Edwards and Grbic-Galic 1994; Phelps and Young 19!and, to a lesser extent, of the xylene isomers and ethylben~Johnston et al. 1996; Gieg et al. 1999! in toluene-degradingmethanogenic systems has been observed.

Lengthy adaptation periods frequently precede the removaBTEX compounds in methanogenic laboratory incubationsrange from approximately 40 to 60 days for toluene~Gieg et al.1999; Phelps and Young 1999! to up to 420 days for benzen~Kazumi et al. 1997!. As noted below in the discussion of MTBdegradation under methanogenic conditions, long lag timelaboratory biodegradation studies may indicate that indigenmicroorganisms are not adapted to the contaminant~s! and maynot be metabolizing these compoundsin situ. On the other handbenzene was biodegraded without a significant lag periodmethanogenic sediments from a petroleum-contaminated aq~Weiner and Lovley 1998b!.

Electron Acceptor and Nutrient RequirementsTwo different approaches can be used to estimate the amounelectron acceptors required to oxidize dissolved BTEX copounds. One method involves the construction of balancedchiometric reactions of electron donor utilization for energy geeration and cell synthesis and is based on thermodynamic coerations~McCarty 1972! ~Table 2!. Importantly, this method accounts for the fact that not all of the electrons derived fromelectron donor~in this case, a BTEX compound! are transferredby growing microorganisms to an electron acceptor for enegeneration. Instead, a fraction of electrons are directed tosynthesis in reactions that consume some of the energy geneby the reduction of the electron acceptor. The relative fractionelectrons consumed in the energy reaction and in cell synthare determined by the free energy changes of the energy (DGR)and synthesis reactions, which in turn are functions of the elecdonor, electron acceptor, carbon source, and nitrogen source

Using benzene as a model compound, balanced stoichiomreactions for energy generation and cell synthesis under aerFe~III !-reducing, denitrifying, sulfate-reducing, and methanogeconditions were determined using the method of McCarty~1972!~Table 2!. For these determinations, it is assumed that benzserves as the electron donor and carbon source, and that NH4

1 isthe nitrogen source. Under methanogenic conditions, the mine



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ization of aromatic compounds is a multistep process and requthe involvement of multiple populations. However, the constrution of a balanced stoichiometric equation under methanogeconditions was simplified by treating the production of methafrom benzene as a single step. The mass requirements for2 ,Fe~III !, nitrate, and sulfate that will theoretically achieve complete oxidation of benzene to CO2 were obtained from the molarelationships defined by the stoichiometric equations and areincluded in Table 2. Mineralization of benzene under methagenic conditions results in a net production of CO2 . Analogousstoichiometric reactions of electron donor utilization for energeneration and cell synthesis and electron acceptor mass reqments can be determined for the other BTEX compounds.

BTEX-biodegrading organisms also utilize electron acceptthrough endogenous decay. However, it is difficult to estimateamount of electron acceptor required to meet the demandendogenous decay without having any information about theof biomass decay. Therefore, the electron acceptor requiremof endogenous decay are neglected in these stoichiometric dminations. It is important to note that the effects of cell syntheand endogenous decay on electron acceptor requirements cbe disregarded simply on the basis of a contaminant plume bat steady state, as has been suggested~Wiedemeier et al. 1999!.Although the steady-state condition may indicate that, overall,growth of BTEX-degrading organisms in the plume is being banced by biomass losses due to decay and other mechanismssynthesis and decay exert a demand for electron acceptors annot cancel each other out.

Reactions catalyzed by oxygenases utilize molecular oxyas an obligatory cosubstrate, which represents an additionaltron acceptor requirement for aerobic BTEX oxidation~Rittmannet al. 1994!. This cosubstrate requirement must be quantifialong with the demand for oxygen as a terminal electron accepThe known metabolic pathways for aerobic BTEX biodegradat~Ellis and Wackett 2001! involve a total of two or threeoxygenase-mediated reactions and thus consume two or tmoles of molecular oxygen per mole BTEX. The oxygen coretant requirement was added to the demand for oxygen as a tenal electron acceptor in the aerobic mineralization of benz~Table 2!.

The second method that can be used to estimate electronceptor requirements simply involves balancing two redox hareactions, i.e., the oxidation of a BTEX compound to CO2 , andthe reduction of an electron acceptor to its reduced form. T


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method neglects the utilization of some of the electron donorsynthesis and oxygen coreactant requirements. Using thisproach, Wiedemeier et al.~1999! constructed balanced reactionfor the mineralization of benzene coupled to the reduction of O2 ,Fe~III !, nitrate, sulfate, and carbon dioxide, and calculatedaverage electron acceptor requirements for oxidation ofBTEX compounds. By comparing the mass ratio of oxygenbenzene in Table 2~1.26 g O2 /g benzene! to the O2 utilizationfactor ~3.14 g O2 /g BTEX! reported by Wiedemeier et al.~1999!,it is clear that, by neglecting cell synthesis, the amount of2

required to achieve complete aerobic oxidation of a BTEX copound is overestimated by a factor of more than two. Howethe magnitude of the errors introduced into estimates of elecacceptor mass requirements by neglecting cell synthesis diishes as the electron acceptors become energetically less fable. This occurs because asDGR values become less negativmicroorganisms have to direct a greater fraction of the electrderived from the electron donor to the electron acceptor forergy generation. Correspondingly, a smaller fraction of electrare consumed for cell synthesis. Similarly, the bioenergetic mopredictions of electron acceptor requirements will approachpredictions of the redox half-reactions under conditions thatnificantly decrease the efficiency of growth@summarized by Mc-Carty ~1972!#.

BTEX-degrading populations, like all bacteria, also requmacro- and micronutrients. However, endogenous sources otrients appear to be adequate to support biodegradation of Bcompounds in most subsurface environments~Wiedemeier et al.1999!.

Biochemistry and Potential Metabolic Indicators of BTEXOxidationA common feature of all the known peripheral and central paways involved in the aerobic biodegradation of the BTEX copounds is that they include reactions mediated by monooxynases and/or dioxygenases~Ellis and Wackett 2001!. Oxygenasesintroduce hydroxyl groups into the BTEX compounds, makithem susceptible to ring cleavage, which is mediated by dioxynases~Heider and Fuchs 1997!. Biodegradation of benzene occuvia initial attack on the aromatic nucleus, forming catechol. Odation of the alkyl-substituted benzenes is initiated via attackther on a side chain or on the aromatic nucleus and generesults in the formation of substituted catechols. Cleavage ofcatechol and substituted catechols formed during the aerobicdation of BTEX compounds occurs via either theortho- or meta-cleavage pathway and leads to the formation of compoundscan be funneled into central metabolic pathways.

Despite the potential importance of anaerobic metabolic pcesses in removing BTEX contamination from the subsurfacompared with the aerobic oxidation of the BTEX compounrelatively little is known about the pathways of BTEX oxidatiounder anaerobic conditions. However, at least two fundamendifferent strategies for initiating anaerobic biodegradationalkylbenzenes have been observed~Spormann and Widdel 2000!.Transformation of methyl-substituted benzenes is initiated bydition to fumarate, an unusual enzymatic reaction that resultthe formation of a new carbon:carbon bond. This type of enmatic activation has been observed in pure cultures of toluedegrading denitrifiers and sulfate-reducing bacteria, a toluedegrading methanogenic enrichment culture~Beller and Edwards2000!, andm-xylene-degrading denitrifiers. Resultant intermeates include benzylsuccinate andE-phenylitaconate in the case otoluene, and methyl homologs of these compounds in the bio



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radation ofm-xylene. In contrast, biodegradation of ethylbenzeunder denitrifying conditions is initiated by dehydrogenationthe benzylic carbon. In the cultures described above, biodegrtion of toluene and ethylbenzene eventually leads to the formaof benzoate, while methylbenzoate is an intermediate in thetabolism ofm-xylene. Benzoate is also a proposed intermediatethe complete oxidation of toluene to carbon dioxide under Fe~III !-reducing conditions~Lovley and Lonergan 1990!.

Anaerobic biodegradation of benzene by a pure culture hasyet been observed. However, mixed culture studies indicatephenol and benzoate may be intermediates in the biodegradof benzene under Fe~III !-reducing, sulfate-reducing, and methnogenic conditions~Caldwell and Suflita 2000!. The conversionof phenol to benzoate by anaerobic mixed cultures via sequecarboxylation and dehydroxylation reactions is well-establish~Heider and Fuchs 1997! and presumably explains the detectioof benzoate in the benzene-degrading enrichment cultures. Hever, production of benzoate via direct carboxylation of benzis conceivable~Caldwell and Suflita 2000!, and an alternativephenol degradation pathway may play a role in the anaerooxidation of benzene in some cases~Weiner and Lovley 1998b!.

As summarized by Beller~2000!, benzoate and methylbenzoates have been observed at a variety of petroleum contaminfield sites, which, along with their relative stability, suggests tthey may be useful indicators of anaerobic BTEX biodegradatHowever, a potential shortcoming of benzoate is that it is a comon intermediate in the anaerobic metabolism of a wide varof substituted aromatic compounds~Heider and Fuchs 1997!.Therefore, its formation is not uniquely indicative of BTEmetabolism. The detection of benzylsuccinate aE-phenylitaconate, or their substituted homologs, can be dnitely related to the anaerobic metabolism of specific alkylbzenes ~Beller 2000!. Further, these compounds are relativestable and are not used for commercial or industrial purposes.utility of these compounds as indicators of BTEX oxidation wexamined by Gieg et al.~1999! in a study evaluating the importance of intrinsic biodegradation at a gas condensacontaminated aquifer. Together with additional chemical andcrobiological evidence, the detection of methylbenzylsucciacids, signature metabolites of anaerobic xylene decay, provstrong evidence that intrinsic biodegradation was occurring atsite. Unfortunately, efforts to quantify benzylsuccinaE-phenylitaconate, and their methyl homologs may be limitedthe lack of commercially available standards.

Biological Mechanisms—MTBE

Two structural features of the MTBE molecule, the ether linkaand methyl branching, are known barriers to the biodegradaof natural and anthropogenic compounds. Therefore, until farecently, MTBE was thought to be recalcitrant~Squillace et al.1998!. However, increasingly, the results of lab and field studindicate that under certain conditions, intrinsic biodegradationMTBE can occur.

Aerobic Degradation of MTBEAerobic biotransformation of MTBE occurs either cometabocally or in the absence of an alternative growth substrate. A vety of pure and mixed cultures are able to utilize short~C3–C5!normal or branched alkanes with a single methyl substitutionprimary growth substrates for the cometabolism of MTBE~Stef-fan et al. 1997; Hyman et al. 1998; Garnier et al. 1999b!. Shortchain alkanes are major components of gasoline and, there


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are likely to be initially present in MTBE-impacted environmenas cocontaminants. However, the alkane compounds foungasoline tend to be relatively immobile due to their low solubties. In contrast, MTBE has a high solubility and a low retardatfactor. As a result, short-chain alkanes are unlikely to be preand support cometabolic transformations throughout MTplumes. Other compounds that have been shown to support ctabolism of MTBE include slightly longer alkanes~Garnier et al.1999a!, alcohols ~Steffan et al. 1997!, camphor ~Steffan et al.1997!, diethyl ether~Garnier et al. 1999a!, and benzene~Koenigs-berg et al. 1999!.

TBA is the most frequently detected intermediate in tcometabolic oxidation of MTBE~Steffan et al. 1997; Hymanet al. 1998!. Two different routes for the cometabolic productioof TBA from MTBE have been proposed~Fig. 2!. In one pathway,hydroxylation of the methoxy group leads to the formation of tunstable hemiacetaltert-butoxymethanol, which spontaneousdismutates to form TBA and formaldehyde. Alternatively, rapoxidation oftert-butoxymethanol totert-butyl formate~TBF!, fol-lowed by biotic and abiotic hydrolysis of TBF to formate anTBA, has been proposed to explain the detection of TBF and Tas intermediates in the cometabolism of MTBE by some orgisms~Hyman et al. 1998!. Additional MTBE cometabolism inter-mediates have been either observed~2-methyl-2-hydroxy-1-propanol and 2-hydroxyisobutyric acid! or proposed~2-propanol,acetone, and hydroxyacetone! ~Steffan et al. 1997!. In somecases, MTBE is not transformed beyond TBA. In other orgisms, cometabolic oxidation of TBA occurs, but is rate-limitintherefore, the intermediate accumulates~Deeb et al. 2000!. Theaccumulation of TBA during aerobic MTBE biotransformationof interest because of its demonstrated carcinogenicity in labmals ~Squillace et al. 1998! and potential use as an indicatoof MTBE bioattenuation~Church et al. 1997!. However, Piveteauet al. ~2001! recently isolated a facultative methylotroph ~strain CIP I-2052! that is able to transform TBA, whichit uses as a sole carbon and energy source, at a relativelyrate. These results, and other observations of TBA degrada@summarized in Deeb et al.~2000!# suggest that complete mine

Fig. 2. Proposed pathways for aerobic biotransformation of MTto TBA



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alization of MTBE is feasible through association of TBAutilizing organisms like strain CIP I-2052 with populations thcometabolically transform MTBE via processes that result inaccumulation of TBA.

CO2 production by some organisms that cometabolize MThas also been observed. Because there are five carbon atomMTBE molecule, the conversion of significantly greater than 20of the radioactivity added as@U-14C#MTBE to 14CO2 indicatesthat the tertiary carbon structure was attacked. For examplestrain ENV425, .60% of @U-14C#-MTBE was converted to14CO2 ~Steffan et al. 1997!. Conversion of 20% or less of thradioactivity added as@U-14C#-MTBE to 14CO2 could result fromthe cleavage and subsequent oxidation of the methoxy megroup alone, unless all of the remaining radioactivity is presen@U-14C#MTBE.

Several isolates~Mo et al. 1997; Hanson et al. 1999! and amixed culture~Salanitro et al. 1994! are able to transform MTBEin the absence of an alternative source of carbon and eneHowever, the results obtained in these studies vary widely wrespect to the percent removal and the extent of MTBE mineization. For example, in a study conducted by Mo et al.~1997!,three bacterial strains were able to transform only 29% ofinitial MTBE dose of 200 ppm over a two-week period and coverted only 8% of the initial radioactivity added a@U-14C#-MTBE to 14CO2 . In contrast, complete removal of a 50mg•L21 dose of MTBE within ten days was achieved withRu-brivivax sp. PM1, and the production of 19%14C-biomass and46% 14CO2 from @U-14C#-MTBE indicates that the tertiary carbon structure was attacked by this organism~Hanson et al. 1999!.Complete removal of MTBE~120 mg•L21! by the mixed cultureBC-1 was also observed~Salanitro et al. 1994!.

Aerobic growth of a pure culture utilizing MTBE as the sosource of carbon and energy has been unequivocally demstrated only for theRubrivivax strain ~Hanson et al. 1999!, al-though increases in biomass associated with MTBE utilizationmixed cultures have also been reported~Park and Cowan 1997!.While these studies indicate that growth on MTBE is feasibleis difficult to speculate about how widely distributed this formmetabolism is and its importance in contaminated environme

TBA was detected as a transient intermediate in the transmation of MTBE by the mixed culture BC-1~Salanitro et al.1994!. Therefore, the biodegradation of MTBE cannot be attruted to either a growth-related or a cometabolic process basethe production of TBA alone. TBA was transformed by BC1 aslower rate than MTBE; however, growth of some culturesMTBE does not result in TBA accumulation~Deeb et al. 2000!.

The studies of aerobic MTBE degradation described abwere conducted with pure or enrichment cultures. Evaluationfield data and experiments involving undefined sediment samsuggest that indigenous organisms at a variety of sites can bigrade MTBE if adequate O2 is available. For example, MTBEremoval was observed in a naturally oxic site where groundwdischarged to a ditch~Landmeyer et al. 2001!. High MTBE levelswere observed in anoxic groundwater unless an artificial sourcO2 was supplied. Significant conversion of@U-14C#MTBE to14CO2 was also observed in surface water sediments taken fsites throughout the United States~Bradley et al. 2001c!. Theamount of14CO2 produced was inversely related to the contentsilt- and clay-sized grains in the sediment. Presumably, coamore permeable grains allowed greater diffusion of oxygen ithe statically incubated sediments.


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Degradation of MTBE under Methanogenic ConditionsThe potential for intrinsic biodegradation of MTBE under anaebic conditions is of interest because the source zone of petrolecontaminated aquifers is typically anaerobic~Fig. 1!. Biodegrada-tion of MTBE under methanogenic conditions has been examiin a variety of systems, including a contaminated groundwaplume~Wilson et al. 2000! and controlled laboratory microcosmcontaining surface and subsurface materials~Mormile et al. 1994;Yeh and Novak 1994; Wilson et al. 2000!. In some of the batchlaboratory experiments, the extent of MTBE transformation wlimited ~Mormile et al. 1994; Bradley et al. 2001b! or requiredthe addition of starch and nutrients~Yeh and Novak 1994!. How-ever, extensive removal of the fuel oxygenate under methagenic conditions has also been observed in the absence of aelectron donors or nutrients~Wilson et al. 2000!. Evidence ofanaerobic MTBE degradation coupled to the reduction of CO2 ,and perhaps other electron acceptors, e.g., Fe~III !, has also beenobtained in aquatic sediment microcosms~Finneran and Lovely2001!. However, transformation of MTBE in each of the labortory batch experiments was preceded by an adaptation perioat least 150 days. Based on the results of these studies, bexperiments designed to evaluate the potential for anaerbiotransformation of MTBE in environmental samples shouhave a duration of at least 200–300 days. In addition, if lengadaptation periods are required in laboratory evaluations, thetential for removal of MTBE from methanogenic regions of cotaminated aquifers may not be realized. As discussed belowpresence of other organic compounds may inhibit MTBE metalism. In a methanogenic source area, other organic compouwill be continuously supplied. Given the high mobility of MTBas compared with other fuel components, MTBE is not likelyreside in methanogenic regions as long as alternative compouTherefore, it is feasible that MTBE will be washed out of methnogenic regions before adaptation can occur.

Increased levels of TBA associated with decreases in MTconcentrations under methanogenic conditions have beenserved in a contaminated groundwater plume~Wilson et al. 2000!and in a laboratory microcosm~Mormile et al. 1994!, suggestingthat TBA is also an intermediate of MTBE biodegradation undmethanogenic conditions.O-demethylation of MTBE has beesuggested as an explanation for the production of TBA in thsystems~Mormile et al. 1994!. @U-14C#TBA was rapidly con-verted to 14CH4 and 14CO2 by a variety of anaerobic TEAPsincluding methanogenesis, without a lag period, in aquatic sment microcosms~Finneran and Lovley 2001!. These results suggest that, if MTBE transformation is initiated under methanogeconditions, complete mineralization may be feasible.

Potential for Degradation of MTBE with Other TerminalElectron AcceptorsBradley et al. ~2001b! observed conversion of;25% of the@U-14C#MTBE radioactivity to14CO2 in NO3

2-amended anoxicsurface water sediments. TBA did not accumulate and appearbe biodegradable. In other studies, MTBE was not transformunder nitrate- or sulfate-reducing conditions~Mormile et al. 1994;Yeh and Novak 1994!. MTBE was rapidly degraded in aquifesediments that were amended with humic substances and Fe~III !,but persisted in unamended sediments~Finneran and Lovley2001!. Presumably, Fe~III !-reducers oxidized MTBE and transferred the electrons to humic substances, which can abiotictransfer electrons to Fe~III ! and are thereby regenerated. In adtion, conversion of a small percentage~;3%! of @U-14C#-MTBE



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to 14CO2 has been observed in microcosms constructed with nrally Fe~III !-reducing aquifer material and groundwater~Landm-eyer et al. 1998!.

Effect of Different Organic Compounds on MTBEDegradationThe presence of a compound that can serve as primary grosubstrate for MTBE-cometabolizing organisms has been showenhance or be essential for MTBE removal via cometabol~Garnier et al. 1999a!. However, short-chainn-alkanes may serveas potential growth substrates and, at the same time, inhgrowth and cometabolic transformation of MTBE through eithspecific ~Garnier et al. 1999b! or nonspecific mechanisms~Ritt-mann et al. 1994!, if present at sufficiently high concentrationGrowth of Rubrivivaxsp. PM1 on either MTBE or benzene habeen demonstrated, and both compounds were used simneously by the pure culture~Deeb et al. 2001!. However, benzeneand MTBE degradation by strain PM1 apparently occur via tindependent and inducible pathways, and MTBE degradationstrain PM1 is inhibited by the BTEX compounds. Cometaboliof MTBE and growth on an alkane has been observed inpresence of BTEX compounds~Steffan et al. 1997!, although in-hibition of MTBE cometabolism by BTEX compounds has albeen noted~Church et al. 2000!. These results suggest that MTBremoval rates may be significantly reduced in the presenceBTEX cocontamination, even if organisms that utilize MTBE assole source of carbon and energy are present. High concentraof endogenous organic matter may also inhibit MTBE removaenvironmental samples under aerobic~Bradley et al. 1999, 2001c!and methanogenic~Yeh and Novak 1994! conditions. Under aero-bic conditions, the observed inhibition of MTBE transformatiomight be attributable to either competitive consumption of eltron acceptors or preferential utilization of natural organic matThe latter might also explain the results observed under methgenic conditions.

Demonstrating Natural Attenuation of PetroleumHydrocarbons and MTBE

Protocols and Procedures

Several protocols and guidance documents on how to demonsnatural attenuation of petroleum hydrocarbons~Wiedemeier et al.1995; ASTM 1998! and MTBE ~Anthony et al. 1999! have beendeveloped. Most protocols have used an independent linesevidence approach to demonstrate that natural attenuation iscurring and at a rate sufficient to be protective of human heand the environment. The focus here is on evidence documenthat the proposed natural attenuation process~es! actually occur~s!in the field. These are the most difficult data to obtain, andmost crucial, because they establish cause and effect by provthe linkage between the evidence of contaminant removal infield and the natural attenuation mechanism~s! ~NRC 1993!. Thistype of evidence may be especially important if the historicontaminant concentration data are inconclusive, or if assessefforts have only recently been initiated~ASTM 1998!. Becauseintrinsic biodegradation contributes most significantly to the naral attenuation of BTEX and MTBE, evidence that this processoccurring is especially important and is emphasized here.


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Table 3. Summary of Example Data Sets from U.S. Field Sites Demonstrating use of Various Combinations of Footprints as Indicators oAttenuation

Site Contaminant~s!Contaminantscontrolled? Footprints

Bemidji, Minn. ~Cozzarelli et al. 1990;Baedecker et al. 1993!

BTEX ~crude oil spill! Yes Loss of O2; production of Fe21, Mn21, andCH4; increase in DIC; increase in alkalinity;production of phenol plus aromatic, alicyclic, andaliphatic acids in anoxic plume; drop inEh

~probe!; DIC enriched in12C under nonmethanogenicconditions; DIC enriched in13C under methanogenicconditions; differential loss of alkylbenzenes

Aquifer near Rocky Point, N.C.~Borden et al. 1995!

BTEX ~gasoline contamination! Yes Loss of O2, NO32 , SO4

22 ; production of Fe21

and CH4; increase in DIC; increase in alkalinity;drop in Eh ~probe!; preferential removal ofcertain BTEX components

Shallow aquifer underlying natural gasproduction site~Gieg et al. 1999!

Gas condensate hydrocarbons~96% w/w C5– C15 compounds,including 18% w/w BTEX!

Yes Loss of O2, NO32 ; SO4

22 , and bioavailable Fe31;production of dissolved sulfides, Fe21 and CH4;increase in alkalinity; detected methylbenzylsuccinicacids, signature metabolites of anaerobic xylenebiodegradation; dissolved H2 levels indicative of ironreduction, sulfate reduction, methanogenesis;10–1,000 fold higher numbers of sulfatereducers and methanogens in contaminated sedimecompared to background

Aquifer in Sampson County, N.C.~Borden et al. 1997!

MTBE ~gasoline and diesel fuel spill! Partially Loss of O2; high NO32 availability; minor Fe21

production; increase in DIC coinciding withBTEX plume;Eh(probe).1200 mV indicatingoxidizing conditions in all wells; mass fluxanalysis indicated MTBE degradation nearcontaminant source, but not in downgradient aquife~supported by laboratory microcosm studies showinlimited MTBE biodegradation and TBA productionnear source but not downgradient!

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‘‘Footprints’’ for BTEX and MTBE

Although the attenuation mechanisms themselves cannot typicbe measured directly, these processes often consume or resproduction of other materials that can be measured i.e., thecesses leave ‘‘footprints’’~NRC 2000!. NRC recommends thacontaminant loss and footprint analyses, coupled with massance analyses and solute fate-and-transport models, form thefor evaluating natural attenuation at a given site. The key meaable footprints of BTEX and MTBE biodegradation are discussin the following paragraphs, including electron acceptors, inganic carbon, alkalinity, metabolic intermediates, redox potenmicrobial numbers and activity, carbon stable isotopes, andratio of nondegradable to degradable substrates. Data setsfield sites demonstrating use of several different combinationfootprints as indicators of intrinsic bioremediation are summrized in Table 3.

Electron Acceptors and Products of Their ReductionConsumption of electron-acceptors correlated with contaminloss can be a key qualitative indicator of intrinsic biodegradat~Chiang et al. 1989; McAllister and Chiang 1994!. Therefore, theconcentrations of electron acceptors, especially O2 , NO3

2, andSO4

22, should be monitored if they are initially present. Fsome TEAPs, the product of reduction of the electron acceptotypically measured rather than the electron acceptor itself.example, Fe~III ! is reduced to Fe~II !, which may be soluble inwater and can be used as an indicator for biodegradation, via



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reduction ~Wiedemeier et al. 1995!. Similarly, the presence omethane at concentrations above background levels may bdicative of methanogenic biodegradation.

Changes in electron acceptors can also be evaluated quatively, based on the stoichiometry of BTEX biodegradation vvarious TEAPs~discussed above!. If the background concentrations of electron acceptors~or their reduction products! areknown, along with their concentrations in the contaminated zothat information can be coupled with the stoichiometric relatioships~Table 2! to provide an indication of the intrinsic capacity othe aquifer system to degrade BTEX via the various TEAPs~Ka-mpbell et al. 1996!. Such calculations are an important partestimating the sustainability of natural attenuation~NRC 2000!.One difficulty with this approach, however, is determining howattribute the observed reduction in electron acceptors to thedation of the different BTEX compounds, or other organic copounds present in the system.

Dissolved Inorganic Carbon„DIC …

As shown in the stoichiometric reactions provided above, the odation of hydrocarbons in groundwater via the various TEAdirectly produces CO2 , which is absorbed by the water, forminmostly HCO3

2 and H2CO3 . As a result, the DIC of the wateshould increase in proportion to the hydrocarbon mineralizatTherefore, estimating the electron acceptor supply rate, ascussed above, and correlating it with increases in DIC, can pvide a quantitative estimate of the rate of petroleum hydrocar


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biodegradation~NRC 2000!. In reality, accurate measurementDIC resulting from biodegradation can be difficult becauseinteractions with the carbonate-buffering system in groundwa~measured as alkalinity! ~ASTM 1998!.

Total Alkalinity „TA …

TA is a parameter that may be influenced by consumptionmicrobial substrates and/or the products of microbial metaboliCO2 produced during biodegradation forms carbonic acid, whmay dissolve carbonate minerals, if present in the aquifer,increase the alkalinity~Chapelle and Bradley 1997!. Therefore,some protocols~Wiedemeier et al. 1995! suggest that increases iTA provide a qualitative indicator of CO2 production and biodeg-radation.

However, in addition to carbonate dissolution, many othchemical and microbially mediated reactions relevant to biorediation can impact the TA and its interpretation~Seagren et al.1998!. Importantly, aerobic respiration alone should not alterTA. In this case, the addition of dissolved CO2 ~or more rigor-ously H2CO3* ! increases the acidity of the system and DIC, bdoes not affect the alkalinity. Alkalinity remains constant becait measures the proton deficiency with respect to the refereproton level, which is H2CO3* or CO2 ~Stumm and Morgan1981!. Similarly, hydrocarbon mineralization under methanogeconditions does not directly affect the TA; however, organic acproduced as intermediates during anaerobic degradation, ascussed above, may contribute to the TA measured~Willey et al.1975!. On the other hand, an increase in TA is expected uniron-, nitrate-, and sulfate-reducing conditions due to proton csumption.

Metabolic IntermediatesOrganic acids are known metabolic intermediates in the anaermicrobial oxidation of petroleum hydrocarbons and, thus, mprovide a footprint of intrinsic biodegradation. In fact, the detetion of a variety of aliphatic and aromatic acids in the anocontaminant plumes at a number of petroleum hydrocarbimpacted sites has been reported~Cozzarelli et al. 1990, 1994Kampbell et al. 1996!, including the ‘‘signature metabolites’’ oanaerobic biodegradation of BTEX compounds, which weretroduced above~Beller et al. 1995; Gieg et al. 1999; Beller 2000!.However, there are some complications with using organic aas a footprint. For example, the transport and fate of organic ain groundwater environments may be complex~Barcelona et al.1993!. Not only are organic acids biologically reactive, but thare also geochemically reactive, impacting processes such aseral dissolution and metal complexation. In addition, significamounts of the organic acids in background and contaminsamples may be associated with the aquifer solids, requiringvanced acid derivatization techniques for quantitative analyFinally, any oxygenated metabolic intermediates produced duaerobic aromatic hydrocarbon degradation are apparently melized rapidly ~Cozzarelli et al. 1990!.

As noted above, the principle daughter product observeding MTBE biodegradation is TBA. TBA may be sufficiently resistant to further degradation to accumulate as an intermed~Church et al. 1997!, which suggests that TBA may be a usefindicator of MTBE biodegradation. However, use of TBA asfootprint of MTBE biodegradation in practice is complicatedthe fact that TBA itself is sometimes used as a fuel additiTherefore, its appearance alone does not provide conclusivedence of MTBE biodegradation~Anthony et al. 1999!. Rather, thedistribution and concentrations of MTBE and TBA should be e



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amined for stoichiometric increases in TBA in association wdecreases in MTBE. In addition, it is difficult to measure loconcentrations of TBA in water~Church et al. 1997!. Finally, todate, there is little conclusive field evidence of TBA resultifrom intrinsic bioremediation of MTBE~Church et al. 1997;Landmeyer et al. 1998; Schirmer and Barker 1998!.

Oxidation-Reduction „Redox… PotentialAs discussed above, the biodegradation of organic contaminin the subsurface typically results in the sequential depletionincreasingly less favorable electron acceptors. Thus, the repotential (Eh) of contaminated groundwater should theoreticadecrease as biodegradation proceeds and may be a useful intor of bioremediation. The classic geochemical indicator of redprocesses is platinum electrode measurement of redox poteHowever, the problems associated with this measurement in cplex systems are well documented~Lovley and Goodwin 1988;Chapelle et al. 1996b!. As a result of these limitations,Eh mea-surements may at best be able to delineate oxic, transitional,anoxic zones~Chapelle et al. 1996b!.

Measurement of H2 concentrations is an alternative indicatof anoxic redox processes~Lovley and Goodwin 1988; Chapelleet al. 1996b!. Based on several studies, Vroblesky and Chap~1994! identified the following ranges of dissolved H2 concentra-tions as being indicative of various TEAPs: denitrification,,0.1nM; iron reduction, 0.1–0.8 nM; sulfate reduction, 1–4 nM; amethanogenesis, 5–25 nM. However, H2 is more reliable if inter-preted in the context of electron acceptor availability andpresence of microbial metabolic byproducts~Chapelle et al.1996b!. In some cases, it may not be possible to discern a patbased on the geochemical indicators and H2 measurements, e.gdue to transport of soluble geochemical indicators, the occurreof different anaerobic processes simultaneously in heterogenmicroenvironments, volumetric averaging of H2 concentrationsduring groundwater sampling, and H2 detection limitations~Gieget al. 1999!. In addition, H2 concentrations may also be affecteby the sampling and pumping methods and the sample-welling material~Chapelle et al. 1997!.

Another product of microbial metabolism that can potentiaprovide an indication of the predominant electron-accepting retions is organic acids~Barcelona et al. 1993; Cozzarelli et a1994!. Using examples from the literature, coupled with a theretical evaluation, Seagren and Becker~1999! concluded that, asthe terminal electron acceptor becomes more reduced, increaconcentrations of aliphatic acids, especially acetate, are obseat petroleum hydrocarbon-contaminated sites. At sites contanated with mono- and polycyclic aromatic hydrocarbons, vations in the composition of the acids are also observed, depenon the dominant TEAP. However, it appears that the relationsbetween microbially mediated redox reactions and the measment of organic acid concentrations, like H2 and platinum elec-trode measurements, may not be unambiguous at some~Vroblesky et al. 1997!.

Microbial Numbers and ActivityCorrelating an increase in the number and activitycontaminant-degrading bacteria with contaminant loss can bkey indication of successful bioremediation, because BTEX mtabolism is often associated with microbial growth~NRC 1993!.For some sites, patterns in both the distribution of microbial nubers and activity have been assessed. For example, at a fuesite near Barrow, Alaska, the number of hydrocarbon-oxidizmicrobes and measured laboratory mineralization potentials


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benzene~at 10°C! were greater in groundwater and soil sampfrom contaminated sites than from reference sites~Braddock andMcCarthy 1996!. However, increases in bacterial numbers abobackground levels may not be detected when biodegradationare low, due to sampling and measurement errors~NRC 1993!.Furthermore, increased numbers of specific types of microbethe contaminated area relative to the background do not prounequivocal evidence of ongoing activity, nor of biodegradatcapabilities ~Gieg et al. 1999!. For example, Salanitro~1993!measured similar BTEX degradation rates, regardless of thesample locations with respect to the hydrocarbon source, althothe highest numbers of BTEX degraders were associatedsamples closest to the source.

Adaptations of native microorganisms leading to the metalism of site contaminants that were either initially persistentwere transformed slowly may also provide evidence of intrinbioremediation~NRC 1993!. To assess adaptation, the rate of cotaminant transformation can be compared in samples fromcontaminated plume and from an adjacent pristine location.nally, correlating increases in bacterial numbers with increaseprotozoa, which prey on bacteria, may also be useful~Madsenet al. 1991!.

Carbon Stable-IsotopesWhen interpretation of changes in DIC concentration or soilCO2 is confounded by abiotic processes~e.g., dissolution of at-mospheric CO2 or carbonate minerals!, carbon stable isotopeanalyses may be applied to the inorganic carbon in liquid andsamples. The ratio of the stable isotopes,13C:12C, is expressed ad13C in parts per thousand~per mil, ‰! ~Stumm and Morgan1981!. A positive change ind13C values reflects a relative enrichment of 13C with respect to12C, and a negative change ind13Cvalues reflects a relative enrichment of12C with respect to13C.

Most petroleum hasd13C values between232 and 221‰~Deines 1980!. As a result, inorganic carbon produced via micrbial oxidation of petroleum is highly enriched in12C as comparedto inorganic carbon from the dissolution of atmospheric C2

(d13C'29 to 27‰! ~Deines 1980!, soil carbonate minerals(d13C'210 to 2‰!, or limestone (d13C'2‰) ~Amundsonet al. 1988!. Indeed, several field studies of in situ petroleuhydrocarbon biodegradation under aerobic conditions haveserved greater enrichment of the soil gas CO2 ~Aggarwal andHinchee 1991; Van de Velde et al. 1995! or groundwater DIC~Baedecker et al. 1993! in 12C than in uncontaminated locations

In addition, isotopic fractionation, i.e., shifting, of carbon istopes also occurs during biologically mediated reactions, becthe bonds formed by light isotopes are more readily broken tthe bonds involving heavy isotopes~Stumm and Morgan 1981!.Thus, the products of the reaction show a preferential enrichmof the lighter isotope, while the remaining source material simtaneously becomes enriched in13C ~Stehmeier et al. 1999!. Vary-ing degrees of fractionation occurs during hydrocarbon biodedation, depending on the TEAP. During aerobic hydrocarbbiodegradation, small amounts of fractionation have generbeen observed, withd13C values of the soil gas CO2 or DICsimilar to that of the petroleum source material~Aggarwal andHinchee 1991; Baedecker et al. 1993; Van de Velde et al. 19Landmeyer et al. 1996!. Somewhat larger fractionation has beobserved during hydrocarbon biodegradation under sulfreducing conditions~Landmeyer et al. 1996!. However, undermethanogenic conditions, extensive fractionation of carbontopes occurs. Metabolically generated CH4 has a very negatived13C ~'2100 to240‰!, while the CO2 produced during metha



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nogenesis has a more positived13C ~'230 to 110‰! ~Whiticaret al. 1986!. Thus, during methanogenic hydrocarbon biodegration, thed13C value of the soil gas CO2 or DIC may be higherthan that of the petroleum source material~Baedecker et al. 1993Landmeyer et al. 1996!.

Compound-specific isotopic analysis may also provide foprints of biodegradation. However, to date, laboratory and firesults are variable for BTEX, and relatively low levels of frationation ~ranging from no change to15.9‰! have generallybeen observed with residual BTEX compounds~Sherwood Lollaret al. 1999; Stehmeier et al. 1999; Ahad et al. 2000!. However,Hunkeler et al.~2001! did observe greater carbon isotopic frationation with residual MTBE~15.1 to 6.9‰!.

Carbon stable isotopic analyses may be complicated by anificant overlap between thed13C of the contaminant hydrocarbons and the natural organic matter~Conrad et al. 1999!. Further-more, the large shifts between thed13C of the substrate and thproduct that occur during some microbially mediated processuch as methanogenesis, can lead to ambiguous results. Tsolve these difficulties, Conrad et al.~1999! suggest also measuring the14C content of the metabolic byproducts; modern levels14C, characteristic of near-surface natural organic matter, are rtively high, while there is no measurable14C in fossil fuels. Forexample, at a site with a plume of aviation-gas derived copounds, CH4 levels in soil gas were high, and thed13C of the CH4

was indicative of methanogenesis via acetate fermenta~'250‰! ~Conrad et al. 1999!. The 14CH4 content, which wasless than 10% of modern levels, confirmed that the CH4 resultedfrom biodegradation of the aviation gas.

Ratio of Degradable to Nondegradable SubstratesIf a mixture of contaminants are present at a site, then a decrwith time in the ratio of biodegradable to nonbiodegradable~i.e.,biomarker! organic contaminants in the field may indicate micrbiological activity~NRC 1993!. An ideal biomarker compound isnonbiodegradable and has physical/chemical properties similathe degradable compound being monitored.

Trimethylbenzene~TMB! isomers have been recommend~Wiedemeier et al. 1995! for use as conservative tracers to esmate the relative portions of observed decreases in BTEX ccentrations that can be attributed to biodegradation or to dission, dilution, and sorption~Wiedemeier et al. 1996!. The TMBisomers have Henry’s constants and soil sorption coefficiesimilar to the BTEX compounds, are generally present at detable levels in contaminated groundwater, and are relativelysistent under anaerobic conditions; however, the degree of recitrance is site-specific. Tetramethylbenzene is another compothat can potentially be used as a conservative tracer~Cozzarelliet al. 1990!.

Intrinsic Biodegradation Rates

Simulation of the fate and transport of BTEX compoundsgroundwater and assessing the efficacy of intrinsic bioremediarequires reliable estimates of biodegradation rate constants.literature is replete with aerobic and anaerobic BTEX biodegdation rate data obtained from laboratory and field studies. Mof these data have been compiled by Wiedemeier et al.~1999!.The methods used to obtain biodegradation rates and the cotions under which these measurements are made may significimpact the values that are obtained. Consequently, it is notprising that, in some cases, reported biodegradation rates foBTEX compounds vary over three orders of magnitude. For


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ample, first-order rate constants for BTEX biodegradation atlected Air Force sites ranged from 0.0002 to 0.08 day21. Becauseof the influence of methodology and site conditions on rate emates, the biodegradation rate values obtained in one studynot be applicable in another system. Thus, caution should beercised in applying rate data from one study to another. In pticular, values obtained from field measurements are typicallyfluenced by site-specific conditions and reflect BTEX oxidatunder a range of TEAPs.

One approach to estimatingin situ biodegradation rates is tomonitor the removal of dissolved hydrocarbons or electron acctors along the groundwater flow path~Barker et al. 1987;Kemblowski et al. 1987; MacIntyre et al. 1993; McAllister anChiang 1994!. First-order kinetic constants have been estimafrom concentration data along the groundwater flowpath bymethods:~1! using a one-dimensional steady-state analyticallution to the advection-dispersion equation~Buscheck and Al-cantar 1995!; or ~2! using a recalcitrant compound~e.g., TMB! inthe dissolved BTEX plume as a conservative tracer~Wiedemeieret al. 1995, 1996!. In an analysis of data from a jet fuel spill, thfirst-order BTEX biodegradation rate constants estimated uthe first method were usually within 20% of the values obtainusing the second method~Wiedemeier et al. 1996!.

A mass-balance approach can also be used for estimatingorder attenuation rate constants due to all attenuation mechan~Chiang et al. 1989; Schirmer et al. 1999!. However, the cost ofthe extensive monitoring network and the relatively complex danalyses that are required preclude its applications at most~McAllister and Chiang 1994!.

A third approach is to estimate biodegradation rates basecontrolled field tracer studies. For example, Thierrin et al.~1993!estimated natural degradation rates with a groundwater traceby injecting deuterated organic compounds and bromide intodissolved hydrocarbon plume downgradient of a leaking undground storage tank. The first-order toluene biodegradationobtained using the tracer test compared well with the modderived field estimates, but was significantly lower than valobtained in laboratory columns.

Laboratory microcosm studies~batch or continuous flow! arethe fourth key approach used to estimate intrinsic biodegradarates based on the disappearance of target compounds~Barkeret al. 1987; Kemblowski et al. 1987; Chiang et al. 1989! or theproduction of 14CO2 from radiolabeled compounds~Chapelleet al. 1996a; Landmeyer et al. 1998!. Such studies have severadvantages, including the capacity to construct a mass balathe ability to distinguish biotic from abiotic removal mechanismby using sterilized treatments, and potential application to a wvariety of hydrologic systems~Madsen 1991; Chapelle et a1996a!. However, the results of laboratory biodegradation evaations using displaced environmental samples are likely toquantitatively, and even qualitatively, different from the resultsthe same determination performedin situ ~Madsen 1991!. There-fore, BTEX rate constants estimated from field-scale studiesmore appropriate for modeling purposes~Wiedemeier et al.1995!, and rates obtained from microcosms are more appropfor demonstrating the potential for biodegradation in the fi~Madsen 1991!. Microcosm studies may be particularly useful fdemonstrating the potential for intrinsic bioremediation of MTBbecause of the difficulties in documenting MTBE biodegradatat the field scale~Anthony et al. 1999!.

First-order BTEX biodegradation rates obtained in field alaboratory studies under aerobic and anaerobic conditions rafrom 0.0002 to 0.0095 day21 for benzene, 0.0017 to 0.2 day21 for



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toluene, 0.0015 to 0.05 day21 for ethylbenzene, and 0.0013 t0.21 day21 for the xylenes~Wiedemeier et al. 1999!. Most trans-formation rates for MTBE have been reported as specific dedation activity ~i.e., mass/cell mass/T!. However, Wilson et al.~2000! did report first-order removal rates for MTBE in anaerobmicrocosms constructed with contaminated aquifer materials.rate of MTBE removal in microcosms supplemented with alkbenzenes was 3.01/year~60.52/year at 95% confidence!, whileremoval in corresponding controls was 0.3960.19/year at 95%confidence. In comparison, removal in microcosms without adalkylbenzenes was 3.560.65/year at 95% confidence~0.360.14/year at 95% confidence in controls!. Also, Hanson et al.~1999! reported removal rates for the pure cultureRubrivivaxsp.PM1, which grows on MTBE, of 0.07, 1.17, and 3.56mg/mL/hfor initial MTBE concentrations of 5, 50, and 500 g/mL, respetively.

Although the first-order kinetics assumption is often advangeous, it is not necessarily appropriate, particularly under cotions with ~1! high and/or toxic substrate concentrations;~2! mul-tiple limiting substrates~e.g., electron donor and acceptor!; and~3! expanding or decreasing microbial populations~Chapelleet al. 1996a!. In some cases, Monod kinetics may be more apppriate for describing BTEX biodegradation in the field~Bekinset al. 1998!. However, to our knowledge, no Monod kinetic prameters have been reported for MTBE and relatively few hbeen reported for the BTEX compounds, especially under anabic conditions. Reported half-saturation coefficient~K! valuesrange from 0.31 to 22.16, 0.044 to 56.74, 11.81, and 0.0007 tomg/L for benzene, toluene, ethylbenzene, and the xylenes, restively. Reported maximum specific growth rate (mmax) valuesrange from 0.784 to 9.3, 0.34 to 10.68, 9, and 3.03 to 11.50 da21

for benzene, toluene, ethylbenzene, and the xylenes, respect~Wiedemeier et al. 1999!. Alternatively, the instantaneous reation model may be used to describe BTEX biodegradation at fisites that are not kinetically limited, i.e., sites with relatively lonhydraulic residence times.

Case StudiesMany case studies demonstrating the natural attenuation of pleum hydrocarbon releases are described in the literature~Barkeret al. 1987; Kemblowski et al. 1987; Chiang et al. 1989; Cozrelli et al. 1990; Baedecker et al. 1993; MacIntyre et al. 19Thierrin et al. 1993; McAllister and Chiang 1994; Wiedemeet al. 1995; Kampbell et al. 1996; Cho et al. 1997!. Indeed, theNRC Committee on Intrinsic Remediation~NRC 2000! rated thecurrent level of understanding and the likelihood of successnatural attenuation for BTEX compounds as ‘‘high.’’

Several extensive compilations of case studies that supporuse of monitored natural attenuation at petroleum-release sitealso available~Hadley and Armstrong 1991; Rice et al. 1995a,Mace et al. 1997; Newell and Connor 1998!. These compilationsindicate that~1! a relatively small proportion of petroleum hydrocarbon releases have affected drinking water wells;~2! benzene isdestroyed near petroleum release source zones by intrinsicdegradation; and~3! groundwater plume lengths at these sitgenerally change slowly and tend to stabilize and/or begincreasing in concentration and length within relatively short [email protected].,,76.2 m~250 ft!, as defined by the 10mg/L benzenecontour# of the hydrocarbon release area.

Some detailed cases studies of BTEX and MTBE naturaltenuation at petroleum release sites are also available. Neveless, the factors controlling MTBE biodegradation in the field a


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still not well understood. Field study findings are variable, raning from observations of low MTBE biodegradation potential,which case reductions in MTBE concentration were primarilytributed to dilution or dispersion~Landmeyer et al. 1998!, to ob-servations of limited MTBE biodegradation under mixed aerobdenitrifying ~Borden et al. 1997! and methanogenic condition~Wilson et al. 2000!. In one study, a mass balance analysis alaboratory experiments suggested that intrinsic biodegradaplayed a major role in MTBE attenuation~Schirmer and Barker1998; Schirmer et al. 1999!; however, field evidence documentinintrinsic biodegradation of MTBE was not obtained~Odencrantz1998!. Correspondingly, the NRC Committee on Intrinsic Remdiation ~NRC 2000! rated the current level of understandingnatural attenuation of MTBE as ‘‘moderate,’’ and the likelihooof success as ‘‘low.’’

Several extensive site data compilations compare MTBEbenzene/BTEX plumes. For example, Happel et al.~1996! ana-lyzed plumes at 30 California leaking underground fuel tank sand found no significant correlation between benzene and MTconcentrations. Further, on a site-by-site basis, benzeneMTBE plumes were not well correlated. MTBE plumes~20 mg/Lcontour! ranged in length from approximately 0.5 to 3.5 times tlength of the respective benzene plume~1 mg/L contour!. In an-other study, Reid et al.~1999! analyzed monitoring data from 8retail gasoline marketing outlets in Florida. The MTBE plumwere generally longer and larger in area than the corresponbenzene plumes, with mean plume length and area, respectof 42.7 m~140 ft! and 1,113.4 m2 ~11,985 ft2! for MTBE and 35.1m ~115 ft! and 735.7 m2 ~7,919 ft2! for benzene, using the 1mg/L contour as the limit. The data analyzed also indicatedbenzene, as well as MTBE, plumes do eventually stabilize: 4of the plumes were expanding, 6.6% of the plumes were staand 89% were decreasing. Surprisingly, estimated first-ordetenuation rates and attenuation rate half-lives for benzeneMTBE were found to be similar.

Summary and Conclusions

Based on this review of the natural attenuation of BTEX aMTBE releases, several observations can be made. First, intrbiodegradation is clearly one of the most important attenuamechanisms contributing to decreases in contaminant concetion and mass for almost all BTEX plumes. Although aerobiodegradation plays an important role, anaerobic processesalso significantly contribute to natural attenuation of BTEX copounds~with the possible exception of benzene at some sites! dueto the abundance of anaerobic electron acceptors as compwith the availability of dissolved oxygen. Second, until fairly rcently, MTBE was considered to be recalcitrant; however,creasingly, laboratory and field results indicate that intrinsic bdegradation of MTBE is feasible under aerobic and anaeroconditions at some sites. Third, field studies have demonstrthat evidence of natural attenuation of BTEX and/or MTBEprovided by key ‘‘footprints,’’ including electron acceptors and/their reduced products, DIC, alkalinity, organic acids, TBA, mcrobial numbers and activity, carbon stable isotopes, and biomkers.

Finally, natural attenuation processes have been observeddocumented at a large number of petroleum-hydrocarbon relsites. At these sites, natural attenuation generally controls thetent of migration of BTEX contaminants away from the sourcethe dissolved plume. On the other hand, the presence of MTB



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gasoline-impacted groundwater may compromise the use of nral attenuation at some sites due to the greater mobility andsistence of MTBE as compared to BTEX in shallow aquifeHowever, some data indicate that MTBE plumes do eventustabilize. Furthermore, field studies indicate that MTBE is biodgradable in shallow aquifers, although the rate of degradamay be slow relative to BTEX compounds. Thus, dispersion adilution may also be important attenuation mechanismsMTBE. Factors contributing to MTBE persistence could inclulengthy adaptation periods for MTBE biodegradation undanaerobic conditions and inhibition of MTBE biodegradationother organic compounds.

Addendum

Since this manuscript was accepted for publication, severalnificant observations involving the anaerobic biodegradationbenzene and MTBE have been made and are in contraststatements made in the text. First, anaerobic biodegradatiobenzene in pure culture has been observed~Coates et al. 2001!.Two Dechloromonasstrains oxidized benzene to CO2 using ni-trate as the terminal electron acceptor. Second, concomitanmoval of MTBE and sulfate reduction have been observed instudies involving sulfate-amended sediment microcosms~Bradleyet al. 2001a; Somsamak et al. 2001!.

References

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Aggarwal, P. K., and Hinchee, R. E.~1991!. ‘‘Monitoring in situ biodeg-radation of hydrocarbons by using stable carbon isotopes.’’Environ.Sci. Technol.,25~6!, 1178–1180.

Ahad, J. M. E., Sherwood Lollar, B., Edwards, E. A., Slater, G. F., aSleep, B. E.~2000!. ‘‘Carbon isotope fractionation during anaerobbiodegradation of toluene: implications for intrinsic bioremediationEnviron. Sci. Technol.,34~5!, 892–896.

Amundson, R. G., Chadwick, O. A., Sowers, J. M., and Doner, H.~1988!. ‘‘Relationship between climate and vegetation and the stacarbon isotope chemistry of soils in the eastern Mojave desert,vada.’’ Quat. Res.,29~3!, 245–254.

Angley, J. T., Brusseau, M. L., Miller, W. L., and Delfino, J. J.~1992!.‘‘Nonequilibrium sorption and aerobic biodegradation of dissolvalkylbenzenes during transport in aquifer material: Column expments and evaluation of a coupled-process model.’’Environ. Sci.Technol.,26~7!, 1404–1410.

Anthony, J. W., et al.~1999!. ‘‘Methodology to evaluate natural attenuation of methyl tertiary-butyl ether.’’Natural attenuation of chlorinatedsolvents, petroleum hydrocarbons, and other organic compounds, B.C. Alleman and A. Leeson, eds., Battelle, Columbus, Ohio, 121–1

ASTM. ~1998!. ‘‘Standard guide for remediation of ground water bnatural attenuation at petroleum release sites.’’ASTM E-1943-98,West Conshohocken, Pa.

Baedecker, M. J., Cozzarelli, I. M., and Eganhouse, R. P.~1993!. ‘‘Crudeoil in a shallow sand and gravel aquifer. III: Biogeochemical reactioand mass balance modeling in anoxic groundwater.’’Appl. Geochem.,8, 569–586.

Barbaro, J. R., Barker, J. F., Lemon, L. A., and Mayfield, C. I.~1992!.‘‘Biotransformation of BTEX under anaerobic, denitrifying condtions: Field and laboratory observations.’’J. Contam. Hydrol.,11~3–4!, 245–272.

Barcelona, M. J., Tomczak, D., Lu, J., and Virkhaus, C.~1993!. ‘‘Frac-tionation and identification of organic matter in natural and fossil-fu


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