Degradation of BTEX by Anaerobic Bacteria: Physiology and Application

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provided by Springer - Publisher Connector Rev Environ Sci Biotechnol (2010) 9:359–385 DOI 10.1007/s11157-010-9219-2

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Degradation of BTEX by anaerobic bacteria: physiology and application

Sander A. B. Weelink • Miriam H. A. van Eekert • Alfons J. M. Stams

Published online: 7 September 2010 The Author(s) 2010. This article is published with open access at Springerlink.com

Abstract Pollution of the environment with aro- these compounds, which is an interesting new alter- matic hydrocarbons, such as benzene, toluene, ethyl- native option for bioremediation. benzene and xylene (so-called BTEX) is often observed. The cleanup of these toxic compounds has Keywords Aromatic hydrocarbons Benzene gained much attention in the last decades. In situ Anaerobic Degradation BTEX Chlorate bioremediation of aromatic hydrocarbons contaminated soils and groundwater by naturally occurring microorganisms or microorganisms that are intro- 1 Introduction duced is possible. Anaerobic bioremediation is an attractive technology as these compounds are often Aromatic compounds are the second most abundant present in the anoxic zones of the environment. The family of organic constituents present in nature after bottleneck in the application of anaerobic techniques carbohydrates. The most important natural sources of is the lack of knowledge about the anaerobic biodeg- aromatic compounds are poorly biodegradable poly- radation of benzene and the bacteria involved in mers such as lignin, condensed tannins and humus anaerobic benzene degradation. Here, we review the (Dagley 1985). Since the start of the industrial existing knowledge on the degradation of benzene and revolution a wide variety of aromatic compounds has other aromatic hydrocarbons by anaerobic bacteria, in also been introduced into the environment through particular the physiology and application, including anthropogenic activity. About 35 million tonnes are results on the (per)chlorate stimulated degradation of produced worldwide annually (http://en.wikipedia. org/wiki/Aromatic). With such huge quantities of these compounds being made, transported and used, it S. A. B. Weelink A. J. M. Stams (&) Laboratory of Microbiology, Wageningen University, is inevitable that a substantial amount will be lost to Dreijenplein 10, 6703 HB Wageningen, The Netherlands the environment. Aromatic compounds are important e-mail: [email protected] constituents of crude oil and oil derivatives. Gasoline is rich in monoaromatic hydrocarbons, such as benzene, M. H. A. van Eekert Lettinga Associates Foundation, Bomenweg 2, toluene, ethylbenzene and xylene (BTEX) (Fig. 1). P.O Box 500, 6700 AM Wageningen, The Netherlands Furthermore, these compounds are used as industrial solvents and they provide the starting materials for the Present Address: production of pharmaceuticals, agrochemicals, poly- S. A. B. Weelink Tauw bv, Handelskade 11, P.O. Box 133, mers, explosives and many other everyday products 7400 AC, Deventer, The Netherlands (Smith 1990). 123 360 Rev Environ Sci Biotechnol (2010) 9:359–385

CH3 CH2CH3 CH3 CH3 CH3

CH3

Fig. 1 Chemical structures of benzene, toluene, ethylbenzene, ortho-xylene, meta-xylene and para-xylene

Accidental spills and industrial discharges have reducing conditions may exist. Further downstream resulted in pollution of the environment with BTEX. and at the fringes of the plume, nitrate and manga- Furthermore, gasoline leakage from underground nese(IV)-reducing conditions prevail (Christensen storage tanks has been identiﬁed as an important et al. 2001). As a result, BTEX contamination is often source of groundwater contamination with BTEX. present in the anaerobic zones of the environment High concentrations of BTEX have been detected in (Lovley 1997). soils, sediments and groundwater. The mobility and Insight into BTEX degradation has led to the toxicity of the BTEX compounds are of major development of biological remediation techniques for concern (several physical–chemical properties of BTEX-contaminated sites. All aromatic compounds BTEX compounds are summarized in Table 1). possess a relative resistance to degradation due to the Compared with other oil hydrocarbons, BTEX are large (negative) resonance energy. This large reso- relatively water-soluble and therefore a plume of nance energy is caused by the stability of the p-electron contamination within the groundwater is formed cloud (Aihara 1992). Before the 1980s, mainly aerobic rapidly (Coates et al. 2002; Chakraborty and Coates BTEX degradation was studied. The last two decades 2004). Benzene is the most hazardous of the BTEX microorganisms that degrade BTEX components in the compounds since it is a known human carcinogen absence of oxygen were also studied. Many bacteria, (leukaemogenic potential) (Badham and Winn 2007). especially Pseudomonas species, have been isolated Toluene and xylene are not carcinogenic, but toluene that can use benzene as sole carbon and energy source can enhance carcinogenesis by other compounds for aerobic growth. Under aerobic conditions, oxygen (Dean 1978). does not only serve as a terminal electron acceptor, but In aquifers contaminated with organic pollutants, it is also used in the initial enzymatic activation of generally a sequence of redox zones has developed as a aromatic compounds. Oxygen is incorporated into the result of organic contamination. Often, near the source aromatic ring and these reactions are catalyzed by of the organic pollutants methanogenic conditions are mono- or dioxygenases (Gibson and Subramanian observed, whereas downstream of the contaminant 1984). Hence, the biochemical strategy for aromatic source zone in the plume sulfate-reducing and iron- hydrocarbon activation under oxic conditions is to

Table 1 Properties of BTEX compounds, according to Van Agteren et al. (1998)

Name Molecular Molecular Density Tm Tb Vapour Aqueous Henry’s Law Log formula weight (kg/l) (C) (C) pressure solubility constant Kow (–) (g/mol) (kPa) (mg/l) (Pa m3/mol)

Benzene C6H6 78.1 0.878 5.5 80.1 10.13 1,780 547 2.13

Toluene C7H8 92.1 0.867 -95 110.8 2.93 515 669 2.65

Ethylbenzene C8H10 106.2 0.867 -95 136.2 0.93 152 588 3.20 ortho-Xylene C8H10 106.2 0.880 -25 144.4 0.67 175 496 2.95 meta-Xylene C8H10 106.2 0.864 -48 139.0 0.80 200 699 3.20 para-Xylene C8H10 106.2 0.860 13 138.4 0.87 198 709 3.18

Density, vapour pressure, aqueous solubility are at 20C. Tm melting point, Tb boiling point 123 Rev Environ Sci Biotechnol (2010) 9:359–385 361 introduce a hydroxyl group (monohydroxylation by 2 Anaerobic BTEX degradation monooxygenase) or hydroxyl groups (dihydroxylation by dioxygenases) into the aromatic ring. The aerobic As BTEX compounds are often present in the anoxic degradation of toluene, ethylbenzene and xylene may zones of the environment, anaerobic bioremediation is involve mono- or dioxygenases, but other pathways an attractive remediation technique. Anaerobic deg- have also been described (Van Agteren et al. 1998). radation of aromatic hydrocarbons was ﬁrst described Aerobic biodegradation of BTEX compounds has been in 1985 (Kuhn et al. 1985). Since then, the degra- studied since the sixties of the last century and has been dation pathways of toluene and ethylbenzene by reviewed several times (Gibson and Subramanian denitrifying and sulfate-reducing microorganisms in 1984; Dagley 1985, 1986; Smith 1990). Under anaer- particular have been characterized. Anaerobic degra- obic conditions, oxygen is not available for the initial dation of BTEX has been reviewed in the past (Schink attack of the ring and therefore other pathways are et al. 1992; Heider et al. 1999; Spormann and Widdel involved in the BTEX degradation. 2000; Phelps and Young 2001; Widdel and Rabus In this review the current knowledge on aerobic 2001; Chakraborty and Coates 2004; Heider 2007; and anaerobic transformation of BTEX compounds is Foght 2008; Fuchs 2008; Carmona et al. 2009). summarized. The emphasis in this paper lies on the Aromatic compounds, such as benzene and toluene, microbial degradation of BTEX compounds under are thermodynamically favorable electron donors for anaerobic conditions. The degradation of BTEX, in growth, because of the high Gibbs free energy change particular benzene, under anaerobic conditions will of the oxidation of these compounds with different be discussed in more detail hereafter. electron acceptors (see Table 2 for stoichiometric

Table 2 Stoichiometric equations and standard free energy changes for benzene (C6H6) oxidations with various electron acceptors with and without biomass (C5H7O2N) formation taken into account

0 Electron acceptor Stoichiometric equation with and without biomass productiona DG0 (kJ/mol)(b) (ox/red)

- - þ ClO3 /Cl C6H6 þ 5ClO3 þ 3H2O ! 6HCO3 þ 5Cl þ 6H -3,813 þ C6H6 þ 1:81ClO3 þ 0:13H2O þ 0:96NH4 ! 1:21HCO3 þ 1:81Cl þ þ0:96C5H7O2N þ 2:17H þ O2/H2O C6H6 þ 7:5O2 þ 3H2O ! 6HCO3 þ 6H -3,173 þ þ C6H6 þ 3:04O2 þ 0:32H2O þ 0:89NH4 ! 1:54HCO3 þ 0:89C5H7O2N þ 2:43H - NO3 /N2 C6H6 þ 6NO3 ! 6HCO3 þ 3N2 -2,978 þ C6H6 þ 2:52NO3 þ 0:87NH4 ! 1:65HCO3 þ 1:26N2 þ 0:87C5H7O2N þ 0:87H2O - - þ NO3 /NO2 C6H6 þ 15NO3 þ 3H2O ! 6HCO3 þ 15NO2 þ 6H -2,061 þ C6H6 þ 7:76NO3 þ 0:83H2O þ 0:72NH4 ! 2:38HCO3 þ 7:76NO2 þ þ0:72C5H7O2N þ 3:11H 3? 2? 3þ 2þ þ Fe /Fe C6H6 þ 30Fe þ 18H2O ! 6HCO3 þ 30Fe þ 36H -3,040 3þ þ 2þ C6H6 þ 12:41Fe þ 6:57H2O þ 0:72NH4 ! 1:60HCO3 þ 12:41Fe þ þ0:88C5H7O2N þ 14:90H 2- 2 þ SO4 /H2S C6H6 þ 3:75SO4 þ 3H2O ! 6HCO3 þ 1:875H2S þ 1:875HS þ 0:375H -186 2 þ C6H6 þ 3:44SO4 þ 2:63H2O þ 0:12NH4 ! 5:38HCO3 þ 1:72H2S þ 1:72HS þ þ0:12C5H7O2N þ 0:34H þ CO2/CH4 C6H6 þ 6:75H2O ! 2:25HCO3 þ 3:75CH4 þ 2:25H -124 þ þ C6H6 þ 6:30H2O þ 0:08NH4 ! 2:04HCO3 þ 3:54CH4 þ 0:08C5H7O2N þ 2:13H a Stoichiometric coefficients, including cell yields, were calculated using method of McCarty (1971) b The data for calculating standard free energy changes (DG0’) are from McCarty (1971), Thauer et al. (1977) and Stumm and Morgan (1981) 123 362 Rev Environ Sci Biotechnol (2010) 9:359–385 equations and standard free energy changes for these aromatics-degrading capacities will be reclassi- benzene oxidations with various electron acceptors). fied in future and separated from the other Azoarcus All studies regarding anaerobic BTEX degradation and Thauera species, which do not have these capac- have indicated that anaerobic benzene degradation is ities (Ku¨hner et al. 2005). The genome of strain EbN1 most difficult and that toluene is one of the aromatic has been sequenced and this bacterium was renamed as compounds, which is relatively easy to degrade Aromatoleum aromaticum (Rabus et al. 2005;Wo¨hl- anaerobically. brand et al. 2007). Recently, an anaerobic betaprote- obacterial strain was described, which uses nitrate, 2.1 Anaerobic toluene biodegradation Fe(III) or Mn(IV) as electron acceptor for growth on toluene. This bacterium (strain G5G6) was described Among the BTEX components, the anaerobic biodeg- as a novel taxon of the Betaproteobacteria, named radation of toluene is probably most extensively Georgfuchsia toluolica (Weelink et al. 2009). studied. Toluene can be biodegraded with nitrate, One of the first reports about anaerobic toluene

Mn(IV), Fe(III), sulfate or CO2 as terminal electron degradation dealt with the degradation of toluene acceptors (Langenhoff et al. 1997b; Chakraborty and coupled to the reduction of Fe(III) by Geobacter Coates 2004 and others). More recently, it was metallireducens GS-15 (Lovley and Lonergan 1990; demonstrated that anaerobic toluene degradation can Lovley et al. 1993). This bacterium belongs to the also be coupled to the reduction of humic substances Deltaproteobacteria and can completely oxidize

(Cervantes et al. 2001), chlorine oxyanions, such as toluene to CO2 coupled to the reduction of Fe(III). perchlorate or chlorate (Chakraborty et al. 2005)or More recently, Geobacter grbiciae, Geobacter sp. arsenate (Liu et al. 2004). Moreover, toluene can be TMJ1 and Desulfitobacterium aromaticivorans were used as a carbon source by anoxygenic phototrophs described which can oxidize toluene with Fe(III) (Zengler et al. 1999). Anaerobic toluene degradation (Coates et al. 2001a; Winderl et al. 2007; Kunapuli has been found in field studies, column studies, et al. 2010). Geobacter metallireducens can also use enrichment cultures and microcosms and in pure nitrate, Mn(IV) or humic substances as the electron cultures (Phelps and Young 2001). Here, pure cultures acceptor (Lovley et al. 1993; Coates et al. 2001a). studies will be discussed in more detail. The betaproteobacterium Georgfuchsia toluolica strain G5G6 also uses nitrate, Fe(III) or Mn(IV) as 2.1.1 Anaerobic toluene-degrading isolates electron acceptor for growth on toluene (Weelink et al. 2009). In general, Geobacter species are often Several bacteria have been isolated, which couple the found to be dominant in the Fe(III)reduction zone degradation of toluene to the reduction of nitrate of environments contaminated with hydrocarbons (Table 3). Most of these bacteria belong either to the (Rooney-Varga et al. 1999; Botton et al. 2007). Azoarcus or Thauera genus, e.g. Thauera aromatica Several bacteria have been described capable of T1 (Evans et al. 1991), Thauera aromatica K172 degrading toluene with sulfate (Table 3), such as (Anders et al. 1995), Azoarcus tolulyticus Tol4 (Fries Desulfobacula toluolica and Desulfotignum toluenicum et al. 1994) and Azoarcus sp. EbN1 (Rabus and Widdel (Rabus et al. 1993; Beller et al. 1996; Harms et al. 1995). Also four Magnetospirillum strains, which 1999b; Meckenstock 1999; Morasch et al. 2004; belong to the Alphaproteobacteria, were described Kuever et al. 2005; Ommedal and Torsvik 2007). that can degrade toluene with nitrate as the electron These sulfate-reducing bacteria all belong to the acceptor (Shinoda et al. 2005). The toluene-oxidizing Deltaproteobacteria. Toluene degradation with other nitrate-reducing Thauera and Azoarcus species are less common electron acceptors has also been reported. facultative anaerobes and are members of the Betapro- Toluene degradation can be coupled to (per)chlo- teobacteria. Most of the these bacteria were isolated rate respiration by Dechloromonas aromatica RCB from anaerobic sludge or (freshwater) sediments (Chakraborty et al. 2005) and to the reduction of (Anders et al. 1995). Several of the Azoarcus and arsenate by strain Y5 (Liu et al. 2004). Finally, toluene Thauera species were originally described as Pseudo- can also be assimilated as a carbon source by the monas species, but were subsequently reclassified. anoxygenic phototroph, Blastochloris sulfoviridis Probably, the Azoarcus and Thauera species that have (Zengler et al. 1999). 123 e nio c itcnl(00 9:359–385 (2010) Biotechnol Sci Environ Rev Table 3 Overview of bacterial strains isolated with benzene, toluene, ethylbenzene and xylene under anaerobic conditions. The electron acceptor that the bacterium uses for growth is indicated in the table Strains Full name or closest relative Growth on References (with % similarity based on 16S rRNA) Benzene Toluene Ethylbenzene Xylenea

- - Strain T Azoarcus sp. strain T NO3 NO3 Dolﬁng et al. (1990) (m) GS-15 Geobacter metallireducens Fe3? c Lovley and Lonergan (1990) - K172 Thauera aromatica NO3 Schocher et al. (1991) - T1 Thauera aromatica T1 NO3 Evans et al. (1991) 2- Tol-2 Desulfobacula toluolica SO4 Rabus et al. (1993) - - 8 Strains Azoarcus tolulyticus Tol4, Azoarcus tolulyticus Td15, NO3 NO3 Fries et al. (1994) Azoarcus toluvorans Td21 (m) - - - 4 Strains:ToN1, PbN1, Aromatoleum aromaticum sp. EbN1 NO3 NO3 NO3 Rabus and Widdel (1995), mXyN1, EbN1 (m)(d) Wo¨hlbrand et al. (2007) - EB1 Azoarcus sp. strain EB1 NO3 Ball et al. (1996) 2- PRTOL1 Desulforhabdus amnigenus (96%) SO4 Beller et al. (1996) - 63 Strains Azoarcus toluclasticus NO3 Fries et al. (1997) - - 14 Strains Azoarcus tolulyticus (97-98%) NO3 NO3 Hess et al. (1997) (m) 2- TRM1 Desulfocapsa thiozymogenes (93%) SO4 Meckenstock (1999) 2- 2- oXyS1 Desulfosarcina ovata (98.7%) SO4 SO4 Harms et al. (1999b) (o) 2- 2- mXyS1 Desulfococcus multivorans (86.9%) SO4 SO4 (m) - pCyN1 Aromatoleum aromaticum sp. EbN1 (100%) NO3 Harms et al. (1999a) ToP1 Blastochloris sulfoviridis phototrophic Zengler et al. (1999) TACP Geobacter grbiciae TACP Fe3? Coates et al. (2001a) -(b) - - - RCB, JJ Dechloromonas aromatica RCB NO3 NO3 NO3 NO3 Chakraborty et al. (2005), (m,o) Coates et al. (2001b) - - Dechloromonas sp. JJ NO3 NO3 - S2 Thauera aminoaromatica S2 NO3 Mechichi et al. (2002) 2- EbS7 Strain mXyS1 (96%) SO4 Kniemeyer et al. (2003) 2- 2- OX39 Desulfotomaculum strain R-acetonA170 (96%) SO4 SO4 Morasch et al. (2004)

123 (m,o) 3- Y5 Desulfosporosinus meridiei (97%) AsO4 Liu et al. (2004) - DNT-1 Thauera aminoaromatica (99%) NO3 Shinoda et al. (2004) 363 364 Rev Environ Sci Biotechnol (2010) 9:359–385

) 2.1.2 Anaerobic toluene degradation pathway 2007 ) ) ) ) The biochemical pathway of anaerobic toluene degra- ) ) 2010 2010 dation has been intensively studied over the last decade 2009 2005 2005

2006 (Heider et al. 1999; Spormann and Widdel 2000; Widdel and Rabus 2001). Especially, Azoarcus strain T, Thauera aromatica K172 and Thauera aromatica T1 were studied (Biegert et al. 1996; Beller and Spormann

Shinoda et al. ( Kuever et al. ( Kasai et al. ( Ommedal and Torsvik ( Weelink et al. ( Kunapuli et al. ( 1997b, 1998; Coschigano et al. 1998; Heider et al. 1998; Leuthner and Heider 1998; Leuthner et al. 1998;

a Beller and Spormann 1999; Krieger et al. 1999). These (o) Kunapuli et al. (

? studies revealed that the ﬁrst step in the catabolism of 3

Fe toluene is the addition of toluene to the double bond of fumarate to form benzylsuccinate by benzylsuccinate synthase (bssABC) (Fig. 2) (Leuthner et al. 1998). Benzylsuccinate is then activated to CoA-thioester by a

) as the electron acceptor succinyl-CoA-dependent CoA-transferase (bbsEF), - - 3 3

(see Validation list no 49, Int J Syst Evol Microbiol, 2004, 44, and benzylsuccinyl-CoA is then converted to succi-

NO nyl-CoA and benzoyl-CoA. Benzoyl-CoA reductase (bcrCABD) initiates the degradation of benzoyl-CoA, which is thereafter further oxidized via reductive (c) - - - 2 2 3 - 3 ? ?

4 4 ring cleavage to carbon dioxide. Benzoyl-CoA has 3 3 Fe SO NO Fe SO been recognized as central intermediate in the anaer- ) and chlorate (ClO - 4 obic degradation of many aromatic compounds

Desulfosarcina cetonica (Harwood et al. 1999). -xylene p

- Although the bss pathway was ﬁrst identiﬁed in 3 Azoarcus and Thauera species growing under nitrate- Growth onBenzene Toluene Ethylbenzene Xylene References reducing conditions, it is now considered as the common mechanism for activation of toluene under various (anaerobic) redox conditions by phylogenetically diverse bacteria. The bss pathway was also found in the toluene-oxidizing Fe(III)-reducing Geobacter metallireducens (Kane et al. 2002), in the sulfate- -xylene and ‘p’ stands for o sp. ToN1 (99%) NO RCB is able to use perchlorate (ClO reducing bacteria, PRTOL1 and Desulfobacula toluo-

AMB-1 (99-100%) NO lica (Beller and Spormann 1997a; Rabus and Heider , but the name has been corrected to 1998) and in the toluene-degrading betaproteobacteri- Azoarcus

(e) um Georgfuchsia toluolica (Weelink et al. 2009). -xylene

m Furthermore, this pathway was demonstrated in the

(99%), toluene-utilizing phototrophic Blastochloris sulfoviridis (Zengler et al. 1999). Bss genes were also identiﬁed -xylene, ‘o’ stands for

m in a methanogenic toluene-degrading culture (Washer Dechloromonas aromatica and Edwards 2007). It is possible that some bacteria ), - 3 use other pathways for the initial toluene conversion, Desulfobacterium cetonicum Desulfosarcina cetonica Geobacter toluenoxydans Georgfuchsia toluolica Desulﬁtobacterium aromaticivorans Azoarcus evansii Desulfotignum toluenicum Magnetospirillum magneticum (with % similarity based on 16S rRNA) such as a pathway involving direct methyl group hydroxylation to benzyl alcohol or hydroxylation to cresol (Frazer et al. 1995; Langenhoff et al. 1997a). continued These pathways have not been studied as thoroughly as the bss pathway. Therefore, involvement of fumarate Besides nitrate (NO Two of four strains able to grow on ‘m’ stands for growth with Nitrate and Mn(IV) can also serve as the electron acceptorOriginally for named toluene oxidation Strain 480 TMJ1 G5G6 UKTL DN11, AN9 H3 4 Strains 370–371 and Validation list no 107, Int J Syst Evol Microbiol, 2006, 56, 1–6) Table 3 Strains Full name or closest relative a b c d e addition cannot be ruled out (Phelps and Young 2001). 123 Rev Environ Sci Biotechnol (2010) 9:359–385 365

COO 2[H] COO

COO COSCoA COSCoA COSCoA COSCoA COO COO HO O O SCoA 2[H] H2O 2[H] HSCoA CH3 COO COO COO COO COO

BssABC BbsEF BbsG BbsH BbsCD BbsAB toluene (R)-benzyl- (R)-benzyl- (E)-phenyl- 2-[hydroxy(phenyl)- Benzoyl- Benzoyl-CoA succinate succinyl-CoA itaconyl-CoA methyl]-succinyl-CoA succinyl-CoA

COSCoA COSCoA

COO COO

Fig. 2 Anaerobic toluene degradation route, according to Kube (R)-benzylsuccinyl-CoA dehydrogenase; BbsH, phenylitaco- et al. (2004). BssABC, benzylsuccinate synthase; BbsEF, nyl-CoA hydratase; BbsCD, 2-[hydroxy(phenyl)methyl]-succi- succinyl-CoA:(R)-benzylsuccinate CoA-transferase; BbsG, nyl-CoA dehydrogenase; BbsAB, benzoylsuccinyl-CoA thiolase

2.2 Anaerobic xylene degradation described (Herrmann et al. 2009). It is noteworthy that the p-xylene-degrading enrichment cultures Although all three xylene isomers can be biodegraded described by Ha¨ner et al. (1995) and Rotaru et al. anaerobically, they have different susceptibilities to (2010) both were very selective; they were unable to anaerobic biodegradation (Foght 2008). Early reports grow with benzene, ethylbenzene or o-xylene. of anaerobic xylene degradation under nitrate-reduc- Several pure cultures have been shown to utilize ing conditions indicated that both p- and m- (but not o-) m-xylene for growth under nitrate- (Dolfing et al. 1990; xylene were degraded (Kuhn et al. 1985), whereas Fries et al. 1994; Rabus and Widdel 1995; Hess et al. Edwards et al. (1992) found p- and o-xylene degrada- 1997) and sulfate-reducing (Harms et al. 1999b) tion (m-xylene was not tested) under sulfate-reducing conditions, whereas only two pure cultures of sulfate conditions. Several subsequent studies demonstrated reducers have been reported to mineralize o-xylene that m-xylene is the most readily degraded isomer in anaerobically, Desulfosarcina sp. strain oXyS1 mixed cultures (Beller et al. 1995). In some cases the (Harms et al. 1999b), renamed as Desulfosarcina ovata presence of m-xylene inhibited (Jahn et al. 2005) (Kuever et al. 2005), and Desulfotomaculum sp. strain concomitant o- and p-xylene degradation (Mecken- OX39 (Morasch et al. 2004). Such isolates often show stock et al. 2004a; Morasch et al. 2004). narrow substrate specificity, with mXyS1 degrading Degradation of o-xylene was demonstrated under only meta- and oXyS1 preferring ortho- substituted denitrifying and sulfate-reducing conditions (Zeyer aromatics (Harms et al. 1999b). No pure cultures et al. 1986; Edwards et al. 1992), and recently the first utilizing p-xylene for growth have yet been reported, iron-reducing enrichment cultures were shown to although Rotaru et al. (2010) described highly enriched oxidize o-xylene (Jahn et al. 2005) and p-xylene Betaproteobacteria growing with p-xylene and nitrate. (Botton and Parsons 2006). Although p-xylene is Phylotypes with 95% 16S rRNA gene sequence identity often reported to be recalcitrant under anoxic condi- to Denitratisoma oestradiolicum dominated the enrich- tions (Rabus and Widdel 1995), it was degraded by ment cultures ([91% of all cells) (Rotaru et al. 2010). enrichment cultures under denitrifying (Ha¨ner et al. The initial steps of anaerobic m-xylene and o-xylene 1995; Rotaru et al. 2010), iron-reducing (Botton and degradation were elucidated in Azoarcus sp. strain T Parsons 2007) and sulfate-reducing conditions (Mor- (Beller and Spormann 1997b; Krieger et al. 1999). asch and Meckenstock 2005; Nakagawa et al. 2008). Research by Krieger et al. (1999)revealedthatm-xylene Recently, sulfate-reducing enrichment cultures that is converted in a series of reactions that are similar to degraded m-, o- and p-xylenes, respectively, were those of fumarate addition in toluene degradation. 123 366 Rev Environ Sci Biotechnol (2010) 9:359–385

Homologs corresponding to toluene fumarate addition group forming benzoylacetate. Benzoylacetate is metabolites have been detected in cultures incubated converted via benzoylacetyl-CoA to benzoyl-CoA with xylenes. For example, 4-methylbenzylsuccinate (Ball et al. 1996), the central intermediate in the and 4-methylphenylitaconate were extracted from an anaerobic catabolism of aromatic compounds. enrichment culture incubated with p-xylene (Morasch A recent study on anaerobic ethylbenzene degra- and Meckenstock 2005), and the expected 2-methylb- dation under sulfate-reducing conditions resulted in enzylsuccinate homolog was detected in cultures the isolation of a novel organism, strain EbS7 co-metabolizing o-xylene (Beller and Spormann (Kniemeyer et al. 2003). This strain is a member of 1997b). In addition, Herrmann et al. (2009)identiﬁed the Deltaproteobacteria, most closely related to strain the benzylsuccinate synthase (bssA) in several xylene- mXyS1, which can anaerobically oxidize toluene and degrading cultures using bssA-targeted primers. m-xylene (Harms et al. 1999b). In contrast to the initial dehydrogenation reaction used by denitrifying 2.3 Anaerobic ethylbenzene degradation ethylbenzene degraders, but similarly to toluene and xylene pathways, activation of ethylbenzene by strain Anaerobic ethylbenzene degradation has been dem- EbS7 is achieved by fumarate addition at the second- onstrated in several nitrate-reducing bacteria and one ary carbon atom of the ethyl group to form 1-phen- sulfate-reducing bacterium (Table 3). Although eth- ylethyl-succinate (Kniemeyer et al. 2003). ylbenzene is chemically very similar to toluene, it is usually degraded via a different pathway. The 2.4 Anaerobic benzene degradation denitrifying bacterial strains EB1, EbN1 (Aromato- leum aromaticum) and PbN1 degrade ethylbenzene to For a long time, benzene was considered to be

CO2. These bacteria were used to elucidate the persistent under anaerobic conditions. Benzene is still anaerobic ethylbenzene degradation pathway (Ball considered as the most recalcitrant of all BTEX et al. 1996; Rabus and Heider 1998; Johnson et al. compounds under anaerobic conditions. A survey of 2001; Rabus et al. 2002). These strains are closely laboratory and ﬁeld studies with groundwater dem- related, and belong to the genus Azoarcus in the onstrated that anaerobic benzene degradation in most Betaproteobacteria. This metabolic pathway (Fig. 3) cases did not occur (Aronson and Howard 1997). included an initial step catalyzed by ethylbenzene Anaerobic benzene degradation has been observed in dehydrogenase, a novel molybdenum/iron-sulfur/ some sediment, microcosm and column studies and heme enzyme (Johnson et al. 2001; Rabus et al. microbial enrichments. Only recently, isolates capa- 2002). It oxidizes the methyl group of ethylbenzene ble of anaerobic benzene degradation have been independently of oxygen, generating (S)-1-phenyl- described (Coates et al. 2001b; Kasai et al. 2006). ethanol as ﬁrst intermediate. Further metabolism of Anaerobic benzene degradation has been extensively (S)-1-phenylethanol proceeds via oxidation to aceto- reviewed (Lovley 2000; Phelps and Young 2001; phenone, which is then carboxylated at the methyl Coates et al. 2002; Chakraborty and Coates 2004).

periplasm cytoplasm

O O O S CoA

HO O O O O S CoA CO2 HS-CoA Acetyl-CoA + + H2O 2 [H] NAD NADH + H (+ATP?) (+ATP?)

ebdABCD ped apc1-5 bal ethylbenzene (S)-1-phenylethanol acetophenone benzoylacetate Benzoylacetyl- Benzoyl-CoA CoA

Fig. 3 Anaerobic metabolism of ethylbenzene, according to Ku¨hner et al. (2005). ebdABC, ethylbenzene dehydrogenase; ped: (S)-1- phenylethanol dehydrogenase; apc1-5, acetophenone carboxylase; bal, benzoylacetate CoA-ligase 123 Rev Environ Sci Biotechnol (2010) 9:359–385 367

2.4.1 Anaerobic benzene degradation under strains (DN11 and AN9) were isolated, which are different redox conditions capable to degrade benzene anaerobically (Coates et al. 2001b; Kasai et al. 2006). Dechloromonas strain Anaerobic benzene degradation with nitrate has been RCB (Dechloromonas aromatica RCB) and strain JJ demonstrated in microcosms (Major et al. 1988), in are phylogenetically closely related (98.1% 16S enrichment cultures (Burland and Edwards 1999; rRNA sequence similarity) and both are members of Ulrich and Edwards 2003) and in pure cultures (Coates the newly described Dechloromonas genus in the et al. 2001b; Kasai et al. 2006). So far anaerobic Betaproteobacteria (Achenbach et al. 2001; Coates benzene degradation with Fe(III) as the electron et al. 2001b). Strain RCB was isolated from river acceptor has not been observed in pure cultures, but sediment with 4-chlorobenzoate as the electron donor only in microcosms (Lovley et al. 1994, 1996a; and chlorate as the electron acceptor. Strain JJ was Anderson et al. 1998; Anderson and Lovley 1999; isolated from lake sediment with anthrahydroquinone Botton and Parsons 2006) and in enrichment cultures 2,6-disulphonate (AQDS, a humic substance analogue) (Rooney-Varga et al. 1999; Botton and Parsons 2007; as the electron donor and nitrate as the electron Kunapuli et al. 2007). Moreover, anaerobic benzene acceptor. Both strains are also able to oxidize benzene degradation with Fe(III) as the electron acceptor was coupled to the reduction of nitrate. Benzene (160 lM) stimulated by humic substances, which serve as was completely degraded to CO2 within 5 days electron shuttle compounds between Fe(III)-reducing (Coates et al. 2001b). So far, subsequent research bacteria and insoluble Fe(III) oxides (Lovley et al. was carried out with strain RCB rather than with strain 1996b). With sulfate as the electron acceptor, anaer- JJ. In addition to benzene, strain RCB is able to use obic benzene degradation has been found in a column toluene, ethylbenzene and m- and o-xylene as the study (Vogt et al. 2007), in microcosms (Edwards and electron donor with nitrate as the electron acceptor. Grbic-Galic 1992; Lovley et al. 1995; Coates et al. Moreover, in addition to nitrate, strain RCB could 1996a, b; Kazumi et al. 1997; Weiner et al. 1998; degrade benzene with perchlorate, chlorate and Anderson and Lovley 2000), and in enrichments oxygen (Chakraborty et al. 2005). The Dechloromonas (Phelps et al. 1996, 1998; Ulrich and Edwards 2003; species, together with the closely related Azospira Musat and Widdel 2007; Herrmann et al. 2008; (previously named Dechlorosoma) species, are con- Kleinsteuber et al. 2008; Laban et al. 2009). Under sidered to represent the predominant (per)chlorate- methanogenic conditions, anaerobic benzene degrada- reducing bacteria in the environment (Coates et al. tion has been demonstrated in microcosms (Kazumi 1999b) and they may be important in the nitrate- et al. 1997; Weiner and Lovley 1998) and in enrich- dependent anaerobic degradation of benzene in the ments (Vogel and Grbic-Galic 1986; Grbic-Galic and environment (Chakraborty and Coates 2004). Vogel 1987; Ulrich and Edwards 2003; Chang et al. Two Azoarcus strains (DN11 and AN9) were 2005). Furthermore, benzene was degraded with isolated, which degrade benzene with nitrate as the Mn(IV) as the electron acceptor in microcosms and electron acceptor (Kasai et al. 2006). Subsequently, sediment columns (Villatoro-Monzon et al. 2003, the degradative capacities of strain DN11 and the 2008). Benzene degradation with (per)chlorate as the potential for its application in bioaugmentation were electron acceptor was shown in pure culture (Coates investigated. Strain DN11 could grow on benzene, et al. 2001b; Weelink et al. 2008) and in column studies toluene, m-xylene and benzoate as the sole carbon (Tan et al. 2006). and energy source under nitrate-reducing conditions, but o- and p-xylene were only cometabolically trans- 2.4.2 Ecology of anaerobic benzene-degrading formed in the presence of toluene. Phenol was not bacteria utilized under anaerobic conditions. Furthermore, strain DN11 could significantly enhance the benzene Little is known about the microorganisms responsible degradation after addition of the strain to laboratory for anaerobic benzene degradation. Most information batches containing benzene-contaminated ground- has been obtained by using molecular approaches. water (Kasai et al. 2007). It was only recently that pure cultures of two Dechlo- Under Fe(III)-reducing conditions, Geobacter spe- romonas strains (RCB and JJ) and two Azoarcus cies have often been associated with anaerobic 123 368 Rev Environ Sci Biotechnol (2010) 9:359–385 benzene degradation. Benzene-degrading Fe(III)- dominated a methanogenic benzene-degrading enrich- reducing sediments and enrichments were dominated ment (Da Silva and Alvarez 2007). It was not proven by bacteria of the family Geobacteraceae (Anderson that this Desulfobacterium was a benzene degrader, et al. 1998; Rooney-Varga et al. 1999; Botton et al. but it was suggested that it either initiates benzene 2007). However, several species in the genus Geob- degradation or is a critical (commensal) partner. Musat acter have the ability to anaerobically degrade and Widdel (2007) demonstrated anaerobic benzene toluene and other aromatic compounds (Lovley et al. degradation by a marine sulfate-reducing enrichment 2004), but none of the Geobacter species tested culture. The dominant phylotype in this enrichment degrades benzene. Recently, a detailed functional and was closely related to a clade of Deltaproteobacteria phylogenetic characterization of a benzene-degrading that includes sulfate-reducing bacteria able to degrade iron-reducing enrichment using stable isotope prob- naphthalene and other aromatic hydrocarbons. Cell ing (SIP) was presented. The authors obtained hybridization with specifically designed 16S rRNA- indications that benzene degradation in the enrich- targeted fluorescent oligonucleotide probes showed ment involved an unusual syntrophy, in which that the retrieved phylotype accounted for more than members of the genus Thermincola within the family 85% of the cells detectable via DAPI staining in the Peptococcaceae primarily oxidize benzene and par- enrichment culture (Musat and Widdel 2007). Liou tially share electrons from benzene with members of et al. (2008) used laboratory incubations of coal-tar the Desulfobulbaceae as syntrophic partners (Kunap- waste-contaminated sediment microbial communities uli et al. 2007). Syntrophic BTX degradation was under relatively controlled physiological conditions also suggested to occur in other iron-reducing (anaerobic with or without sulfate or nitrate additions enrichments (Botton et al. 2007) and in benzene- versus aerobic) to interpret results of a field-based SIP degrading sulfate-reducing consortia (Kleinsteuber assay. By using this SIP approach the authors were able et al. 2008; Herrmann et al. 2010). to associate sets of active benzene-degrading taxa with Phelps et al. (1998) described the molecular char- consumption of particular electron acceptors (nitrate, acterization of a benzene-degrading sulfate-reducing sulfate or mixed aerobic/anaerobic metabolism) and enrichment. This enrichment received benzene as the then qualitatively compare these taxa with those only carbon and energy source for a period of 3 years, retrieved from an uncontrolled in situ field experiment. but repeated attempts to isolate the responsible bacteria 13C-DNA clone libraries revealed a broad diversity of failed. Molecular characterization revealed a diverse taxa involved in benzene metabolism and distinct collection of phylotypes; 16S rRNA genes belonging libraries for each biodegradation treatment. Perhaps to Proteobacteria, Cytophageles and Gram-positive most importantly, in the field SIP experiment the phyla as well as one deeply branching phylotype not clone libraries produced were dominated by Pelomonas closely related to any known bacterium were detected (Betaproteobacteria) sequences similar to those (Phelps et al. 1998). Four clones fell within the found in the anaerobic benzene laboratory experiment Deltaproteobacteria, in the family Desulfobacteria- (Liou et al. 2008). ceae and one of these clones was closely related to a Recently, benzene-degrading sulfate-reducing con- known aromatic hydrocarbon degrader, Desulfobacula sortia were enriched from a benzene-contaminated toluolica. The large diversity of bacteria maintained aquifer (Zeitz field site, Germany). Benzene- over such a long period of time suggests that a mineralizing organisms were enriched under sulfate- consortium of bacteria may be needed to degrade reducing conditions: (1) using in situ microcosms, benzene anaerobically in this enrichment. Members of filled with different solids (sand, lava and Amberlite the family Desulfobacteriaceae have also been found XAD-7), placed for 9 weeks within a groundwater in a methanogenic benzene-degrading enrichment monitoring well located downstream from the source culture. This enrichment was dominated (33% of the zone of the plume, which were subsequently incu- total population) by a clone belonging to the family bated in the laboratory (Herrmann et al. 2008) and (2) Desulfobacteriaceae (Ulrich and Edwards 2003). in columns percolated with benzene-containing In another study, DGGE (Denaturing Gradient Gel anoxic groundwater, filled with either sand or pumice Electrophoresis) analysis demonstrated that a bacte- (Vogt et al. 2007). In control microcosms without rium related to Desulfobacterium sp. clone OR-M2 filling material, benzene was initially degraded, but 123 Rev Environ Sci Biotechnol (2010) 9:359–385 369 the benzene-degrading capacity could not be sus- proposed that, unlike the Cryptanaerobacter/Peloto- tained. Herrmann et al. (2008) suggested that it could maculum phylotype described by Kleinsteuber et al. be favorable to use solids for the in situ enrichment of (2008) and Herrmann et al. (2010), bacteria with anaerobic benzene-degrading bacteria, a strategy that Pelotomaculum related 16S rRNA gene sequences might be useful for the cultivation of bacteria that are oxidize benzene directly with sulfate as electron considered to be hardly or not, cultivable. From the acceptor. In another study, members of the family sand-filled columns, described by Vogt et al. (2007), Desulfobacteraceae have been identified as significant Kleinsteuber et al. (2008) established stable benzene- bacteria in a sulfate-reducing benzene-degrading degrading sulfate-reducing enrichment culture (named enrichment culture, described first by Phelps et al. ZzBs1-4). It was composed of Delta- and Epsilonpro- (1998), using SIP (Oka et al. 2008). teobacteria, Clostridia, Chloroflexi, Actinobacteria and Bacteroidetes. The most prominent phylotype of 2.4.3 Anaerobic benzene degradation pathway the consortium was related to the genus Sulfurovum, followed by Desulfovibrio sp. and the Cryptanaer- Several mechanisms are known to cleave the aro- obacter/Pelotomaculum phylotype. The Cryptanaer- matic ring anaerobically for aromatic compounds obacter/Pelotomaculum phylotype appeared to be with functional groups such as carboxyl or hydroxyl specific for benzene as a growth substrate and might groups. Anaerobic degradation of benzene, however, play a key role in benzene degradation in the consor- is more difficult due to the stability of benzene. The tium. Based on the possible functions of the commu- mechanisms of activation and further degradation of nity members, a functional model for syntrophic benzene are still unknown. The possible initial steps benzene degradation under sulfate-reducing condi- are hydroxylation, carboxylation and methylation, tions was proposed (Kleinsteuber et al. 2008) that was and subsequent transformation to the central aromatic later confirmed by the DNA-SIP experiment described intermediate benzoyl-CoA (Fig. 4), which is further by Herrmann et al. (2010). In this enrichment degraded to CO2. Below, the different pathways that two phyloptypes were dominant: a member of the have been proposed will be discussed in more detail. Cryptanaerobacter/Pelotomaculum group within A. Benzene hydroxylation the Peptococcaceae, and a phylotype belonging to the Epsilonproteobacteria. Herrmann et al. (2010) Studies of Grbic-Galic and Vogel demonstrated hypothesized that the Cryptanaerobacter/Pelotoma- that benzene could be degraded under methanogenic culum phylotype was responsible for the first steps of conditions (Grbic-Galic 1986; Vogel and Grbic-Galic benzene degradation. It was proposed that they form 1986; Grbic-Galic and Vogel 1987). Phenol, cyclo- metabolites like hydrogen, acetate or other low- hexanone, and propionate were detected as puta- molecular mass fermentation products that are used tive intermediates, suggesting initial hydroxylation by the other bacterial partners (such as the Epsilon- of benzene to phenol and subsequent ring reduction proteobacteria phyloptype) and the Methanosaeta- of phenol to cyclohexanone. Experiments with ceae-like Archaea. This would imply that benzene is 18O-labeled water suggested that the oxygen incorpo- mineralized by a consortium consisting of syntrophs, rated into the aromatic ring was derived from water. hydrogenotrophic sulfate reducers and to a minor Weiner and Lovley (1998) found phenol, propionate, extent of aceticlastic methanogens. The function of the and acetate as intermediates in a benzene-degrading Epsilonproteobacteria phyloptype could not be eluci- methanogenic enrichment using 14C-labeled benzene. dated completely, which means that sofar it is not clear Furthermore, benzoate, phenol, o-hydroxybenzoate, how this bacterium degrades the fermentation products and acetate were detected in a benzene-degrading (Herrmann et al. 2010). In another study, different enrichment derived from a mixed inoculum of cow molecular biology techniques showed 95% dominance dung and anaerobic digester sludge (Chaudhuri and of a phylotype that is affiliated to the Gram-positive Wiesmann 1995). In another study using 13C-labeled bacterial genus Pelotomaculum, also found in a benzene, phenol was detected as intermediate in benzene-degrading sulfate-reducing enrichment cul- sulfate-reducing, Fe(III)-reducing and methanogenic ture (Laban et al. 2009). Since there was only one enrichments, and benzoate was detected as interme- dominant organism present, Laban et al. (2009) diate in the sulfate-reducing enrichment (Caldwell 123 370 Rev Environ Sci Biotechnol (2010) 9:359–385

Fig. 4 Possible A mechanisms of the initial O SCoA OH COOH steps of benzene H O 2 [H] CO degradation under 2 2 anaerobic conditions: a benzene hydroxylation, b benzene methylation, c benzene carboxlation OH benzene phenol 4-hydroxybenzoate benzoyl-CoA

B COO O SCoA CH3 CH3-X X fumarate COO

Bss

benzene toluene (R)-benzylsuccinate benzoyl-CoA

C O SCoA

CO2 COOH

benzene benzoate benzoyl-CoA and Suflita 2000). Mass spectral results indicated that phenol to form p-hydroxybenzoate (4-hydroxybenzo- the carboxyl group of the produced benzoate was ate). Several pure cultures degrade phenol anaerobi- 13C-labeled during 13C-benzene degradation. cally via initial carboxylation and this has been It is possible, but difficult to hydrate molecules with observed under nitrate-reducing, sulfate-reducing double bonds that lack an adjacent carbon with a and methanogenic conditions (Heider and Fuchs functional group, such as benzene (Coates et al. 2002). 1997). Anaerobic phenol degradation in Thauera It is not known which enzymes are involved in this aromatica, for instance, starts with the phosphoryla- reaction, but a recently purified molybdenum-con- tion to form phenylphosphate, which is then carbox- taining, iron-sulfur protein that adds a hydroxyl group ylated by phenylphosphate carboxylase, forming to ethylbenzene may provide a model for hydroxyl- p-hydroxybenzoate (Schu¨hle and Fuchs 2004). The ation of benzene (Johnson et al. 2001). In most p-hydroxybenzoate is further activated by a specific prokaryotic systems studied, these oxomolybdenum CoA ligase and the hydroxyl group is reductively enzymes catalyse the transfer of an oxygen atom from removed (Heider and Fuchs 1997). However, in water to the substrate (Coates et al. 2002). Hydrox- studies dealing with anaerobic benzene degradation, ylation of the aromatic ring can also occur by a non- p-hydroxybenzoate has never been detected as inter- enzymatic mechanism involving highly reactive mediate. Evidence for benzene degradation via phenol hydroxyl radicals (HO), which can be formed by a formation and subsequent conversion to benzoate was

Fenton-like reduction of hydrogen peroxide (H2O2)in found in a benzene-degrading methanogenic enrich- the presence of ferrous iron complexes. These radicals ment culture (Ulrich et al. 2005) and in a benzene- can be incorporated into the aromatic ring. However, degrading iron-reducing enrichment (Botton and it is not clear how hydrogen peroxide could be formed Parsons 2007). anaerobically though such a mechanism may be In Dechloromonas aromatica RCB using nitrate as possible at interfaces between aerobic and anaerobic terminal electron acceptor initial hydroxylation of zones (Coates et al. 2002). The conversion of phenol benzene to phenol and subsequent carboxylation to to benzoate probably involves the carboxylation of p-hydroxybenzoate, and loss of the hydroxyl group to 123 Rev Environ Sci Biotechnol (2010) 9:359–385 371 form benzoate (or the CoA derivative, benzoyl-CoA) for anaerobic benzene degradation and that it is similar was found (Chakraborty and Coates 2005) (Fig. 4, to the pathway in strain RCB (Fig. 4, pathway A). pathway A). GC–MS analysis revealed the transient The genome of Dechloromonas aromatica strain formation of phenol and benzoate (Chakraborty and RCB has been sequenced in 2005 (GenBank, http:// Coates 2005). To ensure that the benzene degradation www.ncbi.nlm.nih.gov/, accession no. CP000089) and with nitrate as the electron acceptor in strain RCB was has recently been published (Salinero et al. 2009). not due to a mechanism involving molecular oxygen, Dechloromonas aromatica has a single circular DNA chemical reductants (0.5 mM sodium dithionite or chromosome with a length of 4,501,104 bp and 4,204 0.1 mM sodium ascorbate) were added to active predicted protein coding genes. A function can be benzene-degrading cultures to remove any traces of assigned to 69% of the protein coding genes. Several oxygen. In the presence of the reductants, 14C-benzene putative genes can be found in the genome sequence 14 was still rapidly oxidized to CO2. Formation of that could be involved in benzene metabolism. phenol was detected in cultures amended with sodium Recently, the genome of Dechloromonas aromatica dithionite, however, no phenol was detected in was analyzed. Analysis indicated that gene families benzene-degrading cultures of strain RCB with oxygen that constitute metabolic cycles presumed to create as the electron acceptor. Although the benzene Dechloromonas aromatica’s environmental ‘foot- metabolism was not inhibited, addition of dithionite print’ indicate a high level of diversification between (0.5 mM) retarded the benzene degradation rate and its predicted capabilities and those of its close relatives, subsequent phenol formation. Higher concentrations Aromatoleum aromaticum strain EbN1 and Azoarcus of dithionite ([1 mM) completely inhibited benzene BH72. For example, the benzylsuccinate synthase degradation. However, even at a dithionite concentra- (bssABC) genes (responsible for fumarate addition to tion of 0.5 mM, the phenol degradation rates remained toluene) and genes involved in the central benzoyl- unaffected. It was suggested that dithionite inhibited CoA pathway for monoaromatics are missing in the the initial hydroxylation step of benzene degradation Dechloromonas aromatica genome. However, many in strain RCB rather than the subsequent oxidation of enzymes responsible for aerobic aromatic degradation phenol to CO2 (Chakraborty and Coates 2005). To were found inside the genome (Salinero et al. 2009). investigate the origin of the hydroxyl group of phenol, As D. aromatica was described to degrade benzene 18 experiments with O-labeled H2O were performed. anaerobically with nitrate, it cannot be excluded that However, hardly any incorporation of 18O label into during nitrate reduction oxygen is formed which ini- the hydroxyl group of phenol was found in benzene- tiates oxygenase-dependent benzene degradation as degrading culture of strain RCB and it was concluded described below for chlorate. Oxygen formation linked that H2O is not the source of the hydroxyl group. to denitrification was recently shown for anaerobic Unfortunately, it was not investigated whether the methane-degrading culture (Ettwig et al. 2010). If a hydroxyl group of phenol could originate from nitrate, comparable oxygenic mechanism would exist under 18 - for instance using O-labeled NO3 . On the other nitrate-reducing conditions in strain RCB, the addition hand, it was investigated whether highly reactive of chemical reductants such as dithionite or ascorbate hydroxyl free radicals (HO) were involved in the to the medium had no effect and thus oxidation by hydroxylation of benzene to phenol in strain RCB. oxygenases can not be excluded. They found that hydroxyl free radical scavengers, Azoarcus strain DN11 could not grow on phenol, such as sodium iodide (0.5 mM), propyl iodide indicating that this strain uses a different anaerobic (0.5 mM), 5,5-dimethyl-1-pyrroline-N-oxide, or man- benzene degradative pathway than Dechloromonas nitol (10 mM), inhibited benzene degradation with strain RCB (Kasai et al. 2007). The benzene-degrad- nitrate as the electron acceptor. These results sug- ing sulfate-reducing enrichment described by Musat gested that hydroxyl free radicals play an important and Widdel (2007) did not show metabolic activity role in the benzene ring hydroxylation. Since phenol towards phenol or toluene. Based on this observation and benzoate have been detected in several studies as the authors suggested that benzene degradation by the intermediates of anaerobic benzene degradation under enrichment does not proceed via anaerobic hydroxyl- different redox conditions, Chakraborty et al. (2005) ation to phenol or methylation to toluene (Musat and 13 suggested that a single universal pathway may exist Widdel 2007). Recently, in a study with C6-labeled 123 372 Rev Environ Sci Biotechnol (2010) 9:359–385 benzene as the growth substrate for a benzene- The first direct evidence for benzene methylation degrading iron-reducing enrichment culture, it was was recently found by Ulrich et al. (2005). In a demonstrated that caution should be exercised in nitrate-reducing and methanogenic enrichment cul- interpreting hydroxylated benzene derivatives as ture [ring-13C]-toluene was detected by GC/MS when metabolic intermediates of anaerobic benzene degra- 13C-benzene was added to these cultures. The nature dation (Kunapuli et al. 2008). Phenol was identified as of the methyl group and the subsequent toluene an intermediate, however, it was clearly shown that degradation pathway are not known, but degradation phenol was formed abiotically by autoxidation of via the benzylsuccinate synthase pathway was benzene during the sampling and analysis procedure hypothesized (Ulrich et al. 2005). as a result of exposure to air. This results in the C. Benzene carboxylation production of hydroxyl radicals which readily react with aromatic compounds such as benzene producing Carboxylation of benzene to form benzoate (or phenol and hydroxylation products of phenol. The benzoyl-CoA) has been reported (Caldwell and Suflita authors suggested that autoxidation during sampling 2000; Phelps et al. 2001). In a sulfate-reducing could also be a possible reason for non-detection of enrichment, deuterated benzoate (D5) was detected label in the hydroxyl group of phenol when the when deuterated benzene (D6) was added to the anaerobic benzene-degrading, denitrifying Dechloro- enrichment, but phenol was not detected (Phelps et al. 18 monas strain RCB was incubated with labeled H2 O 2001). The carboxyl group of benzoate was not (Chakraborty and Coates 2005). Chakraborty and labeled when 13C-bicarbonate or 13C-acetic acid were Coates (2005) ruled out water or air as possible added to the enrichment. This is in contrast to the sources of the hydroxyl group, as only 1% label proposed pathway for anaerobic naphthalene and incorporation was observed when the culture was phenanthrene metabolism, where carboxylation of the 18 incubated with H2 O and also phenol production aromatic ring by carbon dioxide (or bicarbonate) is the occurred when dithionite was added to scavenge any initial activation step (Zhang and Young 1997; oxygen present. As water is the only reasonable Meckenstock et al. 2000). The lack of incorporation oxygen source for microorganisms in the absence of of 13C-labeled bicarbonate into the aromatic ring of molecular oxygen, the hydroxyl oxygen had to derive benzene (Phelps et al. 2001) and the suggestion that from air, perhaps by autoxidation during sampling the carboxyl group of benzoate may be derived from (Kunapuli et al. 2008). benzene itself (Caldwell and Suflita 2000) could argue against a direct carboxylation mechanism. B. Benzene methylation However, Kunapuli et al. (2008) demonstrated that Benzene is susceptible to substitution by strong in a benzene-degrading iron-reducing enrichment electrophiles as in Friedel–Crafts alkylation. The culture the carboxyl group of benzoate derived from ? electrophile could be the H3C unit derived from the bicarbonate buffer, indicating a direct carboxyl- 13 S-adenosylmethionine. Biologically mediated meth- ation of benzene. With C6-labeled benzene as the ylation of benzene to toluene has previously been growth substrate, C7-labeled benzoate appeared, observed with human bone marrow incubated with indicating that the carboxyl group of benzoate derived

S-adenosylmethionine (Flesher and Myers 1991). from CO2 that was produced from mineralization Methylation of benzene to toluene is energetically of labeled benzene. This was confirmed by growing favorable using S-adenosylmethionine as the methyl the culture in 13C-bicarbonate-buffered medium donor (Coates et al. 2002). Methylation has also been with unlabeled benzene as the substrate, as the label proposed for the anaerobic activation of another appeared in the carboxyl group of benzoate produced. unsubstituted aromatic hydrocarbon, naphthalene. When using unlabeled buffer and high concentrations This reaction has been found in a sulfate-reducing of 13C-labeled benzene, a significant portion of the enrichment culture and the methyl group might bicarbonate buffer became labeled by the release of 13 possibly be generated from bicarbonate via a HCO3 from complete benzene oxidation. As the reverse CO-dehydrogenase pathway (Safinowski and carboxyl group stems from the bicarbonate buffer, this 13 Meckenstock 2005). eventually leads to the generation of C7-benzoate.

123 Rev Environ Sci Biotechnol (2010) 9:359–385 373

One could speculate that this finding might also fractionation has the ability to identify biodegradation explain why other authors could detect 13C-carboxy- of aromatic hydrocarbons in the field and to distinguish labeled benzoate when applying labeled benzene contaminant mass loss due to biodegradation versus together with unlabeled bicarbonate buffer (Caldwell that due to (abiotic) physicochemical processes. and Suflita 2000). Recently, benzoate was found In studies concerning stable isotope fractionation as intermediate compound in a benzene-degrading of benzene, carbon (12C and 13C) and hydrogen (1H sulfate-reducing enrichment, supporting a direct and 2H) isotope ratios are expressed in the delta carboxylation of benzene as the initial activation notation (d13C and d2H) in per mil (%) units mechanism. However, additional reactions leading to according to the following equation: the formation of benzoate could not be excluded 13 definitely (Laban et al. 2009). d Csample or 2 Rsample Rstandard d Hsample ½&¼ 1; 000 2.4.4 Characterizing and assessing anaerobic Rstandard benzene degradation: isotopic fractionation ð1Þ In this equation, R and R are the Biodegradation is the most important process leading sample standard 13C/12Cor2H/1H-ratio of the sample and an interna- to a decrease in BTEX concentrations in soil and tional standard, respectively. Vienna Pee Dee Bel- groundwater. Therefore, the evaluation of in situ emnite (VPDB) was used as the standard for the BTEX biodegradation is essential for the implemen- analysis of carbon isotope signature and Vienna tation of groundwater management strategies such Standard Mean Ocean Water (VSMOW) was used as as natural attenuation (NA). Traditional methods used the standard for the detection of hydrogen isotope to confirm bioremediation in the field included ratios. For the description of isotope fractionation of monitoring decreases in contaminant concentrations biochemical reactions the Rayleigh equation can be and electron acceptors, and increases in microbial used: (by)products (such as carbon dioxide). This is difficult - e for benzene because of e.g. high background HCO3 1;000 Rt Ct concentrations. However, the challenge is to prove ¼ ð2Þ R C biodegradation in the field, since other processes such 0 0 as volatilization, dispersion, and sorption can cause where Rt, Ct and R0, C0 are the stable isotope ratios loss of the contaminant, and accurate mass balances are and concentrations of a compound at a given point in difficult to obtain (Mancini et al. 2003). In recent years, time and at the beginning of a transformation reaction, stable isotope fractionation analysis has gained atten- respectively. The enrichment factor e [%] provides tion as a tool for characterizing and assessing in situ the link between the changes in stable isotope ratios biodegradation of organic pollutants in contaminated (Rt /R0) and the changes in the concentrations (Ct /C0) aquifers (Meckenstock et al. 2004b; Fischer et al. (Fischer et al. 2007). In recently published laboratory 2007). This concept relies on the fractionation of stable experiments, carbon and hydrogen isotope discrimi- isotopes during the microbial degradation, leading to nation were determined for aerobic and anaerobic an enrichment of heavier isotopes in the residual benzene biodegradation (Hunkeler et al. 2001; fraction of a pollutant. Isotopes of elements such as Mancini et al. 2002; Mancini et al. 2003; Fischer carbon (12C and 13C) and hydrogen (1H and 2H) react at et al. 2008; Mancini et al. 2008; Fischer et al. 2009) slightly different rates during mass-differentiating (see Table 4). reactions. During biodegradation, bonds containing The biodegradation pathway of the aerobic cultures the lighter isotopes are preferentially broken, causing used in the study by Fischer et al. (2008) was known, the remaining contaminant to be enriched in the i.e. mono- or dihydroxylation, and therefore a rela- heavier isotopes compared to the original isotopic tion between enrichment factors and biodegrada- value. A large isotopic fractionation effect (primary tion pathway could be deduced. Their results indicate isotope effect) can be observed if a bond (e.g. C–H that carbon enrichments factors for dihydroxylation bond) containing the element of interest is broken or of benzene are significantly smaller compared formed in the rate-limiting step. Stable isotope to monohydroxylation and anaerobic degradation. 123 374 Rev Environ Sci Biotechnol (2010) 9:359–385

Table 4 Comparison of carbon and hydrogen enrichment factors (eC, eH) in aerobic and anaerobic benzene degradation

Culture Enzymatic pathway eC [%] eH [%] Reference

Burkholderia sp. (aerobic) Unknown -3.5 ± 0.3 -11 ± 2 Hunkeler et al. (2001) Acinetobacter sp. (aerobic) Unknown -1.5 ± 0.1 -13 ± 1 Hunkeler et al. (2001) Rhodococcus opacus strain B-4 Dihydroxylation -1.3 ± 0.2 No enrichment, ±5(a) Fischer et al. (2008) Pseudomonas putida strain ML2 Dihydroxylation -0.7 ± 0.1 No enrichment, ±5(a) Fischer et al. (2008) Ralstonia picketii strain PKO1 Monohydroxylation -1.7 ± 0.2 -11 ± 4 Fischer et al. (2008) Cupriavidus necator ATCC 17697 Monohydroxylation -4.3 ± 0.4 -17 ± 11 Fischer et al. (2008) Alicycliphilus denitriﬁcans strain Monohydroxylation expected -2.6 ± 0.8 -16 ± 4 Fischer et al. (2008) BC (aerobic) Alicycliphilus denitriﬁcans strain Monohydroxylation expected -1.5 ± 0.5 -18 ± 6 Fischer et al. (2008) BC (chlorate) Nitrate-reducing enrichment Unknown, C–H bond cleavage -2.6 ± 0.6 -47 ± 11 Mancini et al. (2008) Swamp NO3-1b expected Nitrate-reducing enrichment Unknown, C–H bond cleavage -2.2 ± 0.4 -35 ± 6 Mancini et al. (2003) Swamp NO3-2b expected Nitrate-reducing enrichment Unknown, C–H bond cleavage -2.8 ± 0.6 -47 ± 4 Mancini et al. (2008) Cart NO3-PW1 expected Nitrate-reducing enrichment Unknown, C–H bond cleavage -1.9 ± 0.7 -31 ± 7 Mancini et al. (2008) Swamp NO3-cons expected Sulfate-reducing enrichment Unknown, C–H bond cleavage -3.6 ± 0.3 -79 ± 4 Mancini et al. (2003) Cart SO4-1a expected

Methanogenic enrichment Cart CH4-1 Unknown -2.1 ± 0.1 -59 ± 4 Mancini et al. (2003), -1.1 ± 0.1 -38 ± 6 Mancini et al. (2008)

Methanogenic enrichment OR CH4-1b Unknown -0.8 ± 0.2 -34 ± 8 Mancini et al. (2008) Sulfate-reducing mixed culture Unknown -1.9 ± 0.3 -59 ± 10 Fischer et al. (2008) a Expected range for eH given by the uncertainty of hydrogen isotope analysis

No systematic differences are given for carbon dihydroxylation reaction occurs, because no hydrogen enrichments factors of monohydroxylation and anaer- bond is broken in the ﬁrst irreversible reaction step of obic degradation. The carbon enrichment factors for the transformation pathway (Hunkeler et al. 2001; monooxygenase reaction are higher as benzene di- Mancini et al. 2003, 2008) or low isotope effect can be hydroxylation, which can be expected when a cleavage expected when an epoxylation is the initial reaction of a C–H bond is involved to some extent as revealed in step (Mitchell et al. 2003). Similarly, the transforma- several studies (Hanzlik et al. 1984; Mitchell et al. tion of benzene by mammalian and methane monoox- 2003). Hydrogen enrichment factors for benzene ygenases exhibited a small or even negligible hydrogen monohydroxylation are higher compared to dihydr- isotope effect which was explained by the absence of oxylation. Hydrogen enrichment factors for benzene C–H bond cleavage in the initial enzymatic degrada- degradation under anaerobic (nitrate-reducing, sulfate- tion step (Hunkeler et al. 2001). reducing and methanogenic) conditions were signiﬁ- Dual parameter plots of carbon and hydrogen cantly higher than for aerobic benzene degradation. isotopic data (Dd2H/Dd13C) demonstrated that these The variability in the hydrogen isotope fractionation plots from each culture were linear, suggesting a for the benzene biodegradation can be explained by consistent reaction mechanism as degradation pro- different initial reaction mechanisms. Carboxylation, ceeded. Methanogenic and sulfate-reducing cultures methylation or hydroxylation of benzene might cause showed consistently higher slopes compared to the cleavage of a C–H bond under anaerobic condition. nitrate-reducing cultures providing evidence for dif- In contrast, aerobic benzene biodegradation should ferent initial reaction mechanisms (Fischer et al. 2008; only lead to secondary hydrogen isotope effects if a Mancini et al. 2008).

123 Rev Environ Sci Biotechnol (2010) 9:359–385 375

3 BTEX biodegradation with (per)chlorate Alpha-, Beta-, Gamma- and Epsilonproteobacteria. Almost all (per)chlorate-reducing species in the From information above it is quite clear that major Betaproteobacteria belong to either the Dechloro- differences exist between BTEX degradation under monas or Azospira (formerly Dechlorosoma) genus. aerobic and anaerobic conditions. Bacteria that (Per)chlorate reducing bacteria have been isolated degrade BTEX in the absence of oxygen do not have from a broad diversity of environments, including the ability to employ degradation pathways with both pristine and contaminated soils and sediments mono- or dioxygenases. Therefore, all the pathways (Coates et al. 1999b). This was remarkable, because in anaerobic bacteria are oxygenase-independent. of the limited natural occurrence of (per)chlorate. However, the situation changes when the bacteria However, these bacteria have diverse metabolic are able to produce oxygen during anaerobic respi- capabilities and this could account for their presence ration. This is the case for bacteria that respire with in environments in which (per)chlorate was not - - perchlorate (ClO4 ) and chlorate (ClO3 ). found. Phenotypic characterization revealed that the (per)chlorate reducing bacteria exhibited a broad 3.1 (Per)chlorate reducing bacteria range of metabolic capabilities, including the oxidation of hydrogen, simple organic acids and alcohols, It has been known for more than 50 years that aromatic hydrocarbons, hexoses, reduced humic sub- microorganisms can reduce oxyanions of chlorine stances, both soluble and insoluble ferrous iron, and such as perchlorate and chlorate under anaerobic hydrogen sulﬁde. All of the known (per)chlorate conditions. The high reduction potential of perchlorate reducing bacteria are facultatively anaerobic or micro- - - 0 - - (ClO4 /Cl E = 1.287 V) and chlorate (ClO3 /Cl aerophilic microorganisms, which is understandable E0 = 1.03 V) makes them ideal electron acceptors for because oxygen is produced during (per)chlorate microbial metabolism. When the reduction of reduction. Some, but not all, (per)chlorate reducing (per)chlorate is coupled to electron-transport phos- bacteria can use nitrate as terminal electron acceptor phorylation, microorganisms are able to grow by for growth. So far, all microorganisms capable of (per)chlorate reduction. Until 1996, little was known perchlorate reduction can also use chlorate as the about the biochemical pathway of (per)chlorate reduc- electron acceptor. However, microorganisms capable tion. Many denitrifying bacteria are able to reduce of chlorate reduction are not always capable of using (per)chlorate, but in general, this reduction is not perchlorate as the electron acceptor (Coates and coupled to growth (Coates and Achenbach 2004). Achenbach 2004). These denitrifying bacteria most probably cannot grow on chlorate because of the accumulation of toxic 3.2 Potential of (per)chlorate-reducing bacteria chlorite during (per)chlorate reduction (Fig. 5). There- in bioremediation fore, bacteria are able to grow by dissimilatory (per)chlorate reduction when: (1) these oxyanions (Per)chlorate reducing bacteria can be used for the can be used as electron acceptor and (2) the toxic treatment of perchlorate- and chlorate-containing waste intermediate chlorite is converted into chloride and streams and groundwater (Logan 1998;Urbansky oxygen. 1998), e.g. the bioremediation of perchlorate-contam- Many dissimilatory (per)chlorate reducing bacteria inated groundwater in a packed bed biological reactor have been isolated and described. A comprehensive (Losi et al. 2002) or the treatment of chlorate and overview is given in literature (Coates and Achen- perchlorate contaminated water using permeable barri- bach 2004). (Per)chlorate reducing bacteria are ers containing vegetable oil (Hunter 2002). The ability phylogenetically diverse and can be found in the of (per)chlorate-reducing bacteria to produce oxygen

Fig. 5 (Per)chlorate 2 [H] H2O 2 [H] H2O reduction scheme including involved enzymes - - - - ClO4 ClO3 ClO2 Cl + O2 perchlorate chlorate chlorite reductase reductase dismutase 123 376 Rev Environ Sci Biotechnol (2010) 9:359–385 can also be applied in other bioremediation processes. in sand columns inoculated with both (aerobic) The dismutation of chlorite by (per)chlorate-reducing toluene-degrading and chlorate-reducing enrichment bacteria in anoxic environments can produce extracel- cultures, which indicated a symbiotic relationship lular O2.ThisO2 can for instance be used by bacteria to between the toluene-degrading bacteria and chlorate- degrade hydrocarbons. Especially, hydrocarbons that reducing bacteria (Logan and Wu 2002). Furthermore, are slowly degraded or persistent under anaerobic the addition of chlorate to a soil column polluted with conditions could be degraded by this mechanism benzene showed removal of benzene coupled with (Coates et al. 1999a; Mehboob et al. 2010). An example chlorate reduction (Tan et al. 2006) and the addition of of such a compound is benzene. chlorate to groundwater at a site in the Netherlands Molecular oxygen can be introduced into the resulted in the complete removal of benzene (Langen- anoxic zones of a contaminated environment by hoff et al. 2009). Recently, the degradation of benzene injection of compressed air or O2 below the water coupled to (per)chlorate reduction was studied in more table (Coates et al. 1999a), but this is a costly and detail in enrichments and pure culture studies (Wee- inefficient process due to the low solubility of oxygen. link et al. 2007, 2008). A benzene-degrading chlorate- Hydrogen peroxide is often used as an additional reducing enrichment culture was obtained with reactor soluble O2 source, but this process has some disad- material, which was originally inoculated with mixed vantages, such as toxicity of hydrogen peroxide to material from a wastewater treatment plant and soil many bacteria. Also solid O2-releasing compounds samples obtained from a location contaminated with can be used, such as magnesium peroxide (MgO2)or benzene (Tan et al. 2004). This stable enrichment calcium peroxide (CaO2). These compounds consist culture degraded benzene at rates similar to reported of powdery material and can also be used as injected aerobic benzene degradation rates, but 20–1,650 times slurries for in situ bioremediation. These O2-releasing faster than reported for anaerobic benzene degradation compounds have some advantages, but the high costs (Weelink et al. 2007). The different partners in the of the O2-releasing compound and the difficulty to enrichment culture were identified as a bacteria evenly distribute the compounds (i.e. the O2) over a closely related to either Alicycliphilus denitrificans large area, for instance a contaminated aquifer, may K601T (Mechichi et al. 2003), or Zoogloea resiniphila be major disadvantages. Therefore, chlorite dismuta- PIV-3A2w, or Mesorhizobium sp. WG or Stenotroph- tion by (per)chlorate-reducing bacteria offers a good omonas acidaminiphila (Costa et al. 2000; Assih et al. alternative strategy to supply extracellular oxygen to 2002; Probian et al. 2003). DGGE analysis of cultures the aerobic hydrocarbon-oxidizing population as was grown with different electron donors and acceptors shown by Rikken et al. (1996). Cell suspensions of indicated that the bacterium related to Alicycliphilus (per)chlorate reducing cells of strain GR-1 showed denitrificans K601 is able to degrade benzene coupled formation of (extracellular) oxygen upon the addition to chlorate reduction. The role of the other bacteria of chlorite (Rikken et al. 1996). could not be conclusively determined. The bacterium The potential use of dissimilatory (per)chlorate related to Mesorhizobium sp. WG could be enriched reducing bacteria for bioremediation (of soils and with benzene and oxygen, but not with acetate and sediments) has already been recognized in previous chlorate, while the bacterium related to Stenotropho- studies (Coates et al. 1999a; Logan and Wu 2002). The monas acidaminophila grew with acetate and chlorate, amendment of (per)chlorate-reducing bacteria and but not with benzene and oxygen. As during chlorate chlorite to an anoxic soil led to complete degradation reduction oxygen is produced, an aerobic benzene of 14C-benzene to 14C-carbon dioxide. Furthermore, degradation pathway is most likely in this enrichment. the addition of chlorite in anoxic soils samples, Furthermore, as one of these bacteria seems to be able inoculated with starved cells of the (per)chlorate to oxidize benzene aerobically, but not able to reduce reducer Dechloromonas agitata strain CKB, showed chlorate and another bacterium seems able to reduce that 14C-benzene was rapidly oxidized to 14C-carbon chlorate but not oxidize benzene, cross-feeding involv- dioxide. This observation further enhances the appli- ing interspecies oxygen transfer is a likely mechanism. cability of these kind or organisms to in situ bioreme- However, thus far attempts to demonstrate interspecies diation (Coates et al. 1999a). In another study, oxygen transfer between the bacteria in the consortium increased rates of toluene degradation were observed were not successful. 123 Rev Environ Sci Biotechnol (2010) 9:359–385 377

From the benzene-degrading chlorate-reducing during chlorate reduction is used in oxygenase enrichment culture, a bacterium (strain BC) was reactions, i.e. benzene conversion to catechol by isolated (Weelink et al. 2008) that is able to degrade two sequential monooxygenase reactions (by benzene benzene coupled to chlorate reduction. Strain BC monooxygenase, BC-BMOa) and catechol conver- grows on benzene and some other aromatic com- sion to 2-hydroxymuconic semialdehyde by catechol pounds with oxygen, or in the absence of oxygen with 2,3-dioxygenase (by catechol-2,3 dioxygenase, chlorate as the electron acceptor. Strain BC is a BC–C23O). Dihydroxylation of benzene to catechol denitrifying bacterium, but it is not able to grow on cannot be ruled out, but genes encoding benzene benzene with nitrate. The closest cultured relative is dioxygenases were not detected with the primers used Alicycliphilus denitrificans K601 (type strain), a in this study. Oxygen formed in the dismutation of cyclohexanol-degrading nitrate-reducing betaproteo- chlorite may not only be used as terminal electron bacterium. The 16S rRNA gene sequences are 99.7% acceptor in strain BC but is also used to attack similar, while genomic DNA was 74.5 ± 3.5% molecules by means of oxygenases. This demon- similar as indicated by DNA-DNA hybridization. strates the existence of aerobic benzene bacterial However, physiological differences are apparent. biodegradation pathways under essentially anaerobic Strain BC is not able to degrade cyclohexanol, while conditions by the concerted action of chlorite dismu- strain K601T lacks the ability to reduce chlorate. tases, providing the metabolic oxygen needed by During chlorate reduction, oxygen is produced by aromatic activating and cleaving oxygenases. Thus, it the dismutation of chlorite. Therefore, it seems likely was demonstrated that aerobic pathways can be that benzene is degraded via an aerobic degradation employed under conditions where no external oxygen pathway in strain BC. An aerobic degradation pathway is supplied (Weelink et al. 2008). would require chlorate reductase, chlorite dismutase Biodegradation of benzene by Alicycliphilus den- and oxygenase enzymes. Chlorate reductase and itrificans strain BC under both oxic and chlorate- chlorite dismutase activities were determined in reducing conditions was also included in the study cell extracts of strain BC. Subsequently, the total by Fischer et al. (2008) concerning carbon and hydro- genomic DNA of strain BC was screened for sequence gen fractionation during benzene biodegradation signatures indicating the presence of genes that can (Table 4). The carbon enrichment factors for strain potentially encode the key enzymes, i.e. chlorite BC were higher than reported for dihydroxylation dismutases, ring activating oxygenases and ring activation of benzene, but corresponded more to cleavage oxygenases. Chlorite dismutase activity reported range of carbon enrichment factors for was detected in cell extracts of strain BC, but no monohydroxylation. The hydrogen enrichment fac- signal was obtained using the primers previously tors observed for strain BC were in the same range as reported as designed for the detection of the group of that observed for the monooxygenase-catalyzed reac- genes assigned to this function (Bender et al. 2004), tions (Table 4). The observed carbon and hydrogen but the genetic information available for this group of enrichment factors strongly indicate that benzene is genes is still very scarce (Weelink et al. 2008). attacked by a monooxygenase under aerobic and Screening of the genomic DNA of strain BC for ring chlorate-reducing conditions in strain BC. These activating oxygenases and ring cleavage oxygenases results confirmed the previously mentioned results resulted in the detection of a putative benzene of the molecular biology experiments with strain BC monooxygenase (BC–BMOa) gene and a putative in which also evidence for benzene degradation by catechol 2,3-dioxygenase (BC–C23O) gene. The monooxygenases was obtained. sequence features and the physiological data support The genome of strain BC has been sequenced by the hypothesis that these sequences encode functional the DOE Joint Genome Institute (JGI) (http://www. and active benzene monooxygenases and catechol 2,3- jgi.doe.gov/genome-projects/). With the available dioxygenases used by strain BC (Weelink et al. 2008). genome sequences, insight into the molecular mech- Based on physiological, genetic and biochemical anism of intracellular oxygen transfer and the regu- experiments, a benzene degradation pathway with lation of chlorate-dependent and chlorate-independent chlorate as the electron acceptor was proposed in growth can be obtained and the suggested pathway strain BC (Fig. 6). In this pathway, oxygen produced for intracellular oxygen transfer can be confirmed. 123 378 Rev Environ Sci Biotechnol (2010) 9:359–385

benzene metabolism:

benzene + 2 [H] + O2 phenol + H2O

phenol + 2 [H] + O2 catechol + H2O

catechol + O2 2-hydroxymuconic semialdehyde (2-hms)

2-hms + 8 H2O6 CO2 + 22 [H] chlorate metabolism: - - 5 ClO3 + 10 [H] 5 ClO2 + 5 H2O - - 5 ClO2 5 Cl + 5 O2

2 O2 + 8 [H] 4 H2O Total:

- - benzene + 5 ClO3 6 CO2 + 5 Cl + 3 H2O

OH H 2[H] 2[H] O2 OH O COOH CO2 BC- BC- BC- BMOa BMOa OH C23O OH O2 H2O O2 H2O benzene phenol catechol 2-hms

2[H] Benzene/phenol monooxygenase (BC-BMOa) Catechol 2,3-dioxygenase (BC-C23O) - - - ClO3 ClO2 Cl +O2 clr cld Chlorate reductase (clr) chlorate chlorite chloride oxygen Chlorite dismutase (cld)

Fig. 6 Proposed benzene degradation pathway with chlorate extradiol (meta-) cleavage of catechol to 2-hydroxymuconic as the electron acceptor in strain BC. A proposed stoichiom- semialdehyde (2-hms) and complete oxidation of 2-hms to etric reactions involved in benzene degradation with chlorate carbon dioxide and reducing equivalents. Chlorate metabolism as the electron acceptor in strain BC. [H] stands for reducing involves the reduction of chlorate to chlorite, dismutation of equivalents. Benzene metabolism involves the hydroxylation chlorite into chloride and oxygen and subsequent reduction of of benzene to phenol, phenol hydroxylation to catechol, oxygen to water

Besides intracellular oxygen transfer, the concept of oxygen production is biologically controlled. The interspecies oxygen transfer deserves further study. conversion of (per)chlorate to chlorite requires reduc- Insight into interspecies oxygen transfer may broaden ing equivalents that are formed in the oxidation of an the degradation possibilities of microbial communities organic compound. Without an electron donor in polluted soils. Introduction of chlorate-reducing (per)chlorate is chemically stable. This is not the case strains in a polluted soil may lead to growth of non- for other oxygen-releasing chemicals, like MgO2 and chlorate-reducing aerobic bacteria, with the capacity to CaO2. (Per)chlorate reduction and subsequent oxygen metabolize a broad spectrum of substrates. formation can also be beneficial for other compounds In situ bioremediation by using naturally occurring that are anaerobically difficult to degrade but that are microorganisms or introduced microorganisms is readily degraded aerobically. Compounds of interest often for several reasons (e.g. financial reasons) a are naphthalene, monochlorobenzene (MCB), vinyl- very attractive option. (Per)chlorate-enhanced biore- chloride (Vc), methyl tertiary butylether (MTBE), mediation may become an important and attractive long and short chain alkanes and a variety of specific process for stimulated in situ anaerobic soil remedi- compounds like aromatic amines. ation. The advantage of (per)chlorate-enhanced bio- When (per)chlorate addition to a contaminated remediation compared to oxygen injection is that aquifer will be applied as bioremediation technique, it (per)chlorate has a higher solubility than oxygen. has to be taken into account that (per)chlorate itself is a Unlike oxygen, which is soluble in the mM-range, pollutant as well. Perchlorate is a solid rocket propel- (per)chlorate is soluble in the M-range. Furthermore, lant, and chlorate is applied as herbicide and defoliant (per)chlorate can be considered as a slow-oxygen and is formed during paper bleaching (Versteegh et al. release compound. The advantage of using (per)chlo- 1993; Coates and Achenbach 2004). However, rate over other oxygen-releasing chemicals is that (per)chlorate is also naturally present in Chilean 123 Rev Environ Sci Biotechnol (2010) 9:359–385 379 nitrate deposits and other mineral deposits. It was can cause contaminant attenuation, and accurate mass found recently that the electric discharge in chloride- balances are difficult to obtain (Mancini et al. 2003). containing aerosols may be a major source of naturally In recent years, stable isotope fractionation analysis occurring perchlorate (Dasgupta et al. 2005). Biolog- has gained attention as a tool for characterizing and ical (per)chlorate reduction may therefore be a rather assessing in situ biodegradation of organic pollutants ancient process. (Per)chlorate-reducing bacteria are in contaminated aquifers. Thus, future studies on known to use a wide variety of electron donors for anaerobic benzene degradation should focus on: growth, which include organic acids (such as acetate • The identification of the bacteria responsible for and lactate) as well as inorganic compounds like anaerobic benzene degradation. hydrogen and sulfide (Logan 1998; Wolterink et al. • The elucidation of the anaerobic benzene degra- 2002; Coates and Achenbach 2004). Because dation pathway. (per)chlorate is a pollutant itself, the application of • The prove and quantification of anaerobic ben- (per)chlorate addition to contaminated aquifers has to zene degradation in the field (in particular stable be performed in close consultation with the competent isotope fractionation). authorities. Remaining (per)chlorate after the organic pollutants have been degraded, could be reduced by the introduction of easily degradable organic com- Acknowledgments This research was financed through grants pounds to the remediated site. As a result, no of the graduate school WIMEK (Wageningen Institute for Environment and Climate Research, http://www.sense.nl) and of environmentally harmful compounds are left after SKB (Dutch Center for Soil Quality Management and Knowl- the bioremediation process. edge Transfer, http://www.skbodem.nl). The research was incorporated in the TRIpartite Approaches toward Soil systems processes (TRIAS) program (http://www.nwo.nl/trias/english/ index.html). The authors would like to thank Anko Fischer and 4 Future perspectives on anaerobic benzene Howard Junca for the pleasant collaboration. degradation Open Access This article is distributed under the terms of the Although anaerobic alkylbenzene (toluene, ethylben- Creative Commons Attribution Noncommercial License which permits any noncommercial use, distribution, and reproduction zen and xylene) degradation has been studied in some in any medium, provided the original author(s) and source are detail, understanding of the anaerobic degradation of credited. benzene is still in its infancy. This is mostly due to the lack of available pure or enrichment cultures that can degrade benzene anaerobically and the extremely References slow growth of the microorganisms in the available enrichment cultures. 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