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This article appeared in a journal published by Elsevier. The attachedcopy is furnished to the author for internal non-commercial researchand education use, including for instruction at the authors institution
and sharing with colleagues.
Other uses, including reproduction and distribution, or selling orlicensing copies, or posting to personal, institutional or third party
websites are prohibited.
In most cases authors are permitted to post their version of thearticle (e.g. in Word or Tex form) to their personal website orinstitutional repository. Authors requiring further information
regarding Elsevier’s archiving and manuscript policies areencouraged to visit:
Chang-Ping Yu a, Rula A. Deeb b, Kung-Hui Chu c,⇑a Key Laboratory of Urban Environment and Health, Institute of Urban Environment, Chinese Academy of Sciences, Xiamen 361021, China b ARCADIS US, Inc., Emeryville, CA, USA c Zachry Department of Civil Engineering, Texas A&M University College Station, TX 77843-3136, USA
h i g h l i g h t s
" Estrogens can be degraded by growth-linked, non-growt h-linked reactions, or microbial-assi sted abiotic reactions." Estrogen-degrading bacteria are phylogenetically diverse in activated sludge." Bacteria, algae, and fungi can degrade 17a-ethinyl estradiol and 17b-estradiol through different pathways.
a r t i c l e i n f o
Article history:Received 11 December 2012 Received in revised form 26 January 2013 Accepted 29 January 2013 Available online 19 March 2013
Steroidal estrogens, widespre ad in the environment, are contaminants of potential concern because expo- sure to these compounds can cause adverse impacts on aquatic life. Intensive research efforts have been undertaken in order to better understand the environmental occurrence of these compounds. In addition to physical/chemical reactions, biological processes – microbial biodegradation of steroidal estrogens –play a vital role in determining the fate and transport of these compounds in built and natural environ- ments. This review summarizes the current state of knowledge on the microbiology of estrogen biodeg- radation. Aerobic and anaerobic estrogen-degrading microorganisms are phylogenetically diverse; they are mainly isolated from soils, activated sludge, dental plaque and intestines. Estrogens can be degraded via growth-linked and non-growth-linked reaction s, as well as through abiotic degradation in the pres- ence of selective microorganisms. Current knowledge on estrogen biodegradation kinetics and pathways is limited. Molecular methods are useful in deciphering estrogen-degrading microbial community and tracking the quantity of known degraders in bioreactors with different operating conditions. Future research efforts aimed at bridging knowledge gaps on estrogen biodegradation are also proposed.
Exposure to environmental estrogens is known to cause adverse health effects on aquatic life (Vos et al., 2000; Sumpter and John- son, 2008 ). Evidence of sexual disruption in fish was reported two decades ago (Tyler et al., 1998 ). Subsequent studies confirmedthat the estrogenicity of wastewater effluent is responsible for the feminization of male fish (Sumpter and Jobling, 1995; Folmar et al.,1996; Cody and Bortone, 1997; Jobling et al., 1998 ). These phe- nomena are not confined in the United Kingdom . In recent years,similar observations were reported by researche rs from the US, Ja- pan, Korea and Denmark (Hashimoto et al., 2000; Bjerregaard et al.,2006; Vajda et al., 2008; Li et al., 2009 ). The concern of the estrog- enicity of the treated wastewa ter has quickly prompted intensive research to better understand the removal of estrogen by drinking water and wastewa ter treatment plants (WWTPs) (Johnson and Sumpter, 2001; Vulliet et al., 2011 ), by constructed wetlands (Shappell et al., 2006; Song et al., 2009 ), and by swine manure,poultry litter, and dairy waste disposal systems (Fine et al., 2003;Hutchins et al., 2007; Zheng et al., 2007 ). Furthermore, numerous studies were undertaken to investigate the fate and transport ofestrogens in the natural environm ent, including soils (Fan et al.,2007; Xuan et al., 2008; Dubroca et al., 2009 ), river water (Jürgenset al., 2002; Matsuoka et al., 2005 ), groundwate r and sediment from aquifers (Ying et al., 2003 ), seawater and marine sediment (Ying and Kookana, 2003 ).
Biodegradat ion has been reported as the major removal mech- anism that affects the fate and transport of estrogenic compounds in built and natural environments (Johnson and Sumpter, 2001;Writer et al., 2012 ). Bacterial biodegra dation of estrogens by soil isolates was reported in the 1960s (Coombe et al., 1966 ) and bya number of intestina l and oral microorgan isms in the 1980s (Jar-venpaa et al., 1980; Kornman and Loesche, 1982 ). Until recently,due to the concern of environmental estrogens, intensive research efforts have focused on identification and isolation of estrogen- degrading microorganism s from the environment. These research efforts resulted in numerous publications over the past decade on the biodegradat ion of estrogenic compound s. Several reviews focus on the advanced oxidation, fate, and transport of estrogens (Khanal et al., 2006; Esplugas et al., 2007; Koh et al., 2008; Liu et al., 2009b,c; Combalbert and Hernandez-R aquet, 2010; Racz and Goel, 2010 ). This review is, thus, devoted to the review of re- cent findings regarding the biotransformat ion of estrogens by pro- karyotic and eukaryot ic microorgan isms. In this review, the microbiology of estrogen biodegradat ion is summarized , including degradative strains, estrogen-de grading microbial communi ty, bio- transformat ion mechanis ms, biodegradat ion kinetics and pathways.
2. Estrogens
Estrogens, naturally or synthetically produced, are steroidal hormones. They regulate a wide range of important biological
functions in human and animals. Estrone (E1), 17b-estradiol (E2),and estriol (E3) are produced naturally by all vertebrates, including humans. Among the natural estrogens, E2 is responsible for the developmen t of the female reproductive system in humans. E2 isalso considered as the most potent estrogen when compared toE1 and E3. 17a-ethinyl estradiol (EE2) is a synthetic estrogen that is commonly used for contraception. All of these steroidal estro- gens, as shown in Fig. 1, contain one aromatic ring (ring A), two cyclohex ane rings (rings B and C), and a cyclopentane ring (ringD). (Please see Combalbert and Hernandez-R aquet (2010) for the physical and chemical propertie s of estrogens .) Ring A has a high affinity for human estrogen receptors.
Once these estrogens are metaboli zed in the body, their metab- olites and/or conjugated estrogens (inactive forms of estrogens) are excreted from the body mainly in urine and feces, which enter WWTP systems (Liu et al., 2009c ). During WWTP processes , the conjugat ed estrogens are quickly reverted to free estrogens (i.e.,active forms) through enzymatic reactions carried out by wastewa -ter microorgan isms. Estrogens in wastewa ter are mainly removed by two means during biological treatment: sorption to biosolids and biodegradat ion by microorgan isms present in treatment units.The estrogen-loade d biosolids may be used for land applicati ons,which may then become a long-term source of environmental estrogens (Dubroca et al., 2009 ).
Recent studies show that both heterotrop hic and autotrophic wastewa ter microorgan isms can degrade estrogens, and that some of these microorgan isms can convert estrogens into non-estro genic compound s (Yoshimoto et al., 2004; Yu et al., 2007 ). Nevertheles s,the removal of estrogens by conventional WWTPs is incomplete.Subseque ntly, discharges of treated WWTP effluents into the envi- ronment could be a source of estrogens (Kirk et al., 2002; Yu and Chu, 2009 ).
3. Degradatio n mechani sms of estrogens
Studies monitoring the performance of WWTPs report varied percentages of estrogen removals, ranging from 19% to 94% for E1, 76% to 92% for E2, and 83% to 87% for EE2 (Baronti et al.,2000). These variations are due in part to the differences in biolog- ical (fixed film or suspended growth) and other processes, operat- ing conditions, geological locations of WWTPs, as well as the influent concentr ations of estrogens (Ternes et al., 1999; Baronti et al., 2000; Matsui et al., 2000; Vader et al., 2000; Kirk et al.,2002; Andersen et al., 2003; Drewes et al., 2005; Koh et al.,2008; Yu and Chu, 2009; Liu et al., 2009a ). Solids retention time (SRT) and hydraulic retention time (HRT) are two important oper- ating parameters of a biological treatment system. Long SRTs and HRTs have been reported to improve estrogen removal by activated sludge systems (Koh et al., 2008 ). A long SRT would allow for the growth of slow-growing microorganism s, such as ammonia-o xi- dizing bacteria which are also known to degrade estrogens, and alonger contact time for estrogens sorption to biosolids. When there is no sludge recycling in an activated sludge system, the value of
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Fig. 1. Structures and acronyms for estrogens; E1, estrone; E2, 17b-estradiol; E3, estriol; EE2, 17a-ethnyl estradiol.
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HRT is the same as the SRT. A long HRT can increase the contact time for estrogen sorption to biosolids and thus enhance estrogen removal. While both sorption and degradat ion contribute to the overall estrogen removal by WWTPs, biodegradat ion is considered as a favorable and major mechanism for estrogen removal during WWTP processes . This hypothesis was strongly supported by astudy showing that more than 70% of 14C-labeled E2 was converted into 14C-labeled CO2 by nitrifying activated sludge in 24 h (Laytonet al., 2000 ).
Microorgani sms can degrade steroidal hormone using two pos- sible degradation mechanism s: growth-l inked (metabolic) and non-growth -linked (cometabolic) (Yu et al., 2007 ). For growth- linked degradation , microorgan isms utilize steroidal hormones asenergy and/or carbon source for microbial growth. Four wastewa- ter bacterial isolates, Novosphingobi um tardaugens ARI-1 (hereafterreferred as ARI-1) (Fujii et al., 2002 ), Sphingomona s strain KC8 (hereafter referred as KC8) (Yu et al., 2007 ), Sphingobacterium sp.JCR5 (hereafter referred as JCR5) (Ren et al., 2007 ), and Pseudom o-nas aeruginosa TJ1 (Zeng et al., 2009b ), are reported to degrade E2via growth-linked reactions. Among these four strains, KC8 can also grow on E1 and testosterone as sole carbon sources.
When a cometab olic transformat ion reaction is involved , bacte- ria use their existing enzymes to degrade steroidal hormones. Ascometabolic degradation yields no carbon or energy benefits tothe degradat ive microorgan isms, a primary growth substrate isneeded for sustainable bacterial growth and/or required for induc- ing the expression of degradat ive enzymes . Previous studies show that E2 and EE2 could be cometabolicall y degraded by heterotro- phic bacteria (Yu et al., 2007; Pauwels et al., 2008 ) and by ammo- nia-oxidizing bacteria and nitrifying activated sludge (Yi and Harper, 2007; Skotnicka-Pita k et al., 2009 ). As the concentratio nsof estrogens in wastewater are in the range of ng L�1, the cometa- bolic degradation might be an important removal mechanism for removing these trace estrogens in wastewa ter.
A third mechanis m for E2 and EE2 transformat ion, occurring inthe presence of selected microorgan isms, was abiotic nitration oroxidation, In the presence of a high ammonia concentratio n, anammonia-o xidizing bacterium Nitrosomonas europaea could trans- form EE2 into nitro-estrogen s (Gaulke et al., 2008; Sabirova et al.,2008). This nitration reaction is not biotic; rather, it is an indirect,abiotic reaction between the biogenic nitrite produced by the bac- terium and steroidal estrogens. The nitration reaction occurs favor- ably under high nitrite concentr ations and low pH conditions.Similar to the nitration of EE2 mediated by ammonia-oxid izing bacteria, an indirect EE2 degradation was mediated by manga- nese-oxidiz ing bacteria (Sabirova et al., 2008 ). Sabirova et al.(2008) reported that manganese-ox idizing bacteria Leptothrix dis- cophora and Pseudomona s putida strains can oxidize Mn2+ to form Mn oxides, which are then used to facilitate the oxidative cleavage of EE2.
4. Aerobic biodegradation of estrogens
Aerobic estrogen-degra ding bacteria have been isolated from various built and natural environm ents, including activated sludge,compost, soils, sandy aquifers, and the Baltic Sea (Table 1). These isolates are widespread among different phyla: Proteobacteria ,Actinobacte ria, Bacteroidetes, and Firmicutes. Below is a brief sum- mary of these isolates.
4.1. Isolates from activated sludge
Publications reporting wastewater isolates capable of degrading estrogens increased dramatically in recent years. The strain ARI-1 was the first E2-utilizi ng wastewa ter bacterium isolated from acti-
vated sludge in Japan (Fujii et al., 2002 ). The strain ARI-1 can also degrade E1, but not EE2. In addition, its degradat ion ability toward E1 could not be maintain ed after growing on complex nutrient medium without E2 for 7 d (Roh and Chu, 2010 ). Novosphin gobium sp. strain JEM-1 (hereafter referred as JEM-1) is a wastewater bac- terial isolate similar to ARI-1. The strain JEM-1 can degrade E1, E2,and EE2 and was found present in two activated sludge systems. Inaddition, higher numbers of strain JEM-1 were found to be corre- sponding to higher E1 removals in activated sludge (Hashimotoet al., 2010 ).
Two isolated Rhodococcus strains, Rhodococcus zopfii and Rhodo-coccus equi , showed an ability to degrade natural and synthetic estrogens rapidly into non-estrogenic compounds (Yoshimotoet al., 2004 ). In a later study using the same Rhodococcus speciespurchase d from American Type Culture Collection, the researchers were unable to observe the excellent degradation ability toward EE2 (O’Grady et al., 2009 ). Achromobac ter xylosoxid ans and Ralsto-nia picketii , identified from an enrichment culture from activated sludge, were suggested to be responsib le for the degradation ofE1, E2, E3, and 16a-hydroxyestrone (Weber et al., 2005 ).
Yu et al. (2007) reported fourteen phylogenetical ly diverse E2- degrading wastewater isolates (strains KC1 through KC14). Among the fourteen isolates, five are Sphingomonas strains; two are Amin-obactor strains; two are Flavorobacterium strains; and one each ofBrevundi mmonas strain, Microbacter ia strain, Nocardio idas strain,Rhodococc us strain, and Escherichia strain. These strains showed three different estrogen degradat ion patterns: A through C. Eleven out of the fourteen strains showed the degradation pattern Awhere E2 was rapidly converted to a final product, E1. Two other strains (KC6 and KC7) slowly degraded E2 and E1 over time (deg-radation pattern B). Only one strain, Sphingom onas strain KC8,complete ly degraded E2 and E1 into non-estrogenic compound s(degradation pattern C). In addition, strain KC8 can grow on E1,E2, and testoster one as a sole carbon source (Roh and Chu, 2010 ).None of these 14 strains can degrade EE2. An EE2-degrad ing bacte- rium JCR5, isolated from activated sludge, showed an ability togrow on E1, E2, E3, EE2, pyrene, phenanthre ne, methylbenzen e,and dimethylbenze ne (Ren et al., 2007 ).
Similar to strains ARI-1, JEM-1, and KC8, Pseudomonas aerugin- osa TJ1 could grow on E2 as a sole carbon source (Zeng et al.,2009b). In 2010, more than ten wastewa ter estrogen-de grading bacterial isolates were reported in literature. For example, Kurisuet al. (2010) reported five estrogen-de grading bacteria within the genera Rhodococcus and Sphingomonas. Jiang et al. (2010) reportedfive E2-degra ding strains, belonging to the genus Bacillus, two ofwhich were able to degrade E1. An E2-degrading bacterium, Brev-undimon as diminuta , was isolated from an enrichment culture from activated sludge with estrogen and acetonitrile as substrates (Mul-ler et al., 2010 ).
4.2. Isolates from soils, compost, sandy aquifers, and seas
A few estrogen-uti lizing and degrading bacteria have been iso- lated from various environments. An early study in 1980s reported that an Alcaligenes strain, a soil bacterial isolate, could grow on E2or testosterone as a sole carbon source (Payne and Talalay, 1985 ).From compost, six bacterial isolates were able to metabolize E1, E2and E3, and to cometabolize EE2 (Pauwels et al., 2008 ). These six isolates belong to a-, b- and c-Proteobact eria. Three estrogen- degrading cultures Acinetobacter strain LHJ1, Agromyces strainLHJ3, and Sphingomonas strain CYH were isolated from a sandy aquifer (Ke et al., 2007 ). Recent studies reported that two marine microbial isolates, strains H5 and S19-1, could tolerate 4.1% NaCl and degrade testosterone and E2 (Zhang et al., 2011; Sang et al.,2012). These two marine strains were isolated from the Baltic
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Table 1List of aerobic microorganisms capable of degrading or utilizing steroidal hormones.
Phylogenetic affiliation Degradation ability and mechanism Source of isolates References
Alpha-proteobacteria Aminobacter aminovorans KC7 Degradation of E1, E2 Activated sludge Yu et al. (2007)Aminobacter sp. KC6 Degradation of E1, E2 Activated sludge Yu et al. (2007)Brevundimonas diminuta I Conversion of E2 to E1 Activated sludge Muller et al. (2010)Brevundimmonas vesicularies KC12 Conversion of E2 to E1 Activated sludge Yu et al. (2007)Novosphingobium sp. strain JEM-1 Degradation of E1, E2, EE2 Activated sludge Hashimoto et al. (2010)Novosphingobium tardaugens ARI-1 Metabolism of E1, E2, E3 Activated sludge Fujii et al. (2002)Phyllobacterium myrsinacearum BP1 Degradation of E1, E2, E3; cometabolism of
EE2 in the presence of E1, E2, E3Compost Pauwels et al. (2008)
Sphingomonas sp. CYH Degradation of E1, E2 under both aerobic and anoxic conditions
Artificial sandy aquifer Ke et al. (2007)
Sphingomonas sp. KC8 Metabolism of E2, E1, testosterone Activated sludge Yu et al. (2007) and Roh and Chu (2010)
Sphingomonas sp. KC9 Conversion of E2 to E1 Activated sludge Yu et al. (2007)Sphingomonas sp. KC10 Conversion of E2 to E1 Activated sludge Yu et al. (2007)Sphingomonas sp. KC11 Conversion of E2 to E1 Activated sludge Yu et al. (2007)Sphingomonas sp. KC14 Conversion of E2 to E1 Activated sludge Yu et al. (2007)Sphingomonas sp. ED8 Metabolism of E2, E1 Soil samples from
agricultural fieldsKurisu et al. (2010)
Sphingomonas sp. ED9 Metabolism of E2, E1 Soil samples from agricultural fields
Kurisu et al. (2010)
Beta-proteobacteria Achromobacter xylosoxidans Metabolism of E1, E2 Activated sludge Weber et al. (2005)Alcaligenes sp. Metabolism of E2, testosterone Soil Payne and Talalay (1985)Alcaligenes faecalis Conversion of E2 to E1 and vice versa Intestinal microorganisms Jarvenpaa et al. (1980)Leptothrix discophora (LMG 8142) Production of biogenic Mn oxides to oxidize
EE2 Belgian coordinated collections ofmicroorganisms
Sabirova et al. (2008)
Nitrosomonas europaea ATCC 19718 Cometabolism and Nitration of EE2 ATCC Gaulke et al. (2008) and Skotnicka-Pitak et al.(2009)
Ralstonia pickettii BP2 Degradation of E1, E2, E3; cometabolism ofEE2 in the presence of E1, E2, E3
Compost Pauwels et al. (2008)
Ralstonia sp. Metabolism of E1, E2 Activated sludge Weber et al. (2005)
Gamma- proteobacteria
Acinetobacter sp. LHJ1 Conversion of E2 to E1 Artificial sandy aquifer Ke et al. (2007)Acinetobacter sp. BP8 Degradation of E1, E2, E3; cometabolism of
EE2 in the presence of E1, E2, E3Compost Pauwels et al. (2008)
Acinetobacter sp. BP10 Degradation of E1, E2, E3; cometabolism ofEE2 in the presence of E1, E2, E3
Compost Pauwels et al. (2008)
Buttiauxella Metabolism of E2 and Testosterone Baltic Sea Zhang et al. (2011)Escherichia coli KC13 Conversion of E2 to E1 Activated sludge Yu et al. (2007)Pseudomonas aeruginosa Conversion of E2 to E1 and vice versa Intestinal microorganisms Jarvenpaa et al. (1980)Pseudomonas aeruginosa BP3 Degradation of E1, E2, E3; cometabolism of
EE2 in the presence of E1, E2, E3Compost Pauwels et al. (2008)
Pseudomonas aeruginosa TJ1 Metabolism of E2 Activated sludge Zeng et al. (2009b)Pseudomonas putida MnB1 (LMG 2321)
Production of biogenic Mn oxides to oxidize EE2
Belgian coordinated collections ofmicroorganisms
Sabirova et al. (2008)
Pseudomonas putida MnB6 (LMG 2322)
Production of biogenic Mn oxides to oxidize EE2
Belgian coordinated collections ofmicroorganisms
Sabirova et al. (2008)
Pseudomonas putida MnB29 (LMG 2323)
Production of biogenic Mn oxides to oxidize EE2
Belgian coordinated collections ofmicroorganisms
Sabirova et al. (2008)
Pseudomonas sp. BP7 Degradation of E1, E2, E3; cometabolism ofEE2 in the presence of E1, E2, E3
Compost Pauwels et al. (2008)
Vibrio sp. H5 Metbolism of E2 and Testosterone Baltic Sea Sang et al. (2012)
Actinobacteria Agromyces sp. LHJ3 Degradation of E2 and E3 with formation ofE1 under aerobic condition; degradation ofE2 with formation of E1 under anoxic condition
Artificial sandy aquifer Ke et al. (2007)
Mycobacterium smegmatis Conversion of E2 to E1 and vice versa ;conversion of 16a-hydroxyestrone to E3
Intestinal microorganisms Jarvenpaa et al. (1980)
Microbacteria testaceum KC5 Conversion of E2 to E1 Activated sludge Yu et al. (2007)Nocardioides simplex KC3 Conversion of E2 to E1 Activated sludge Yu et al. (2007)Rhodococcus equi Y50155 Metabolism of E1, E2, E3, EE2 Activated sludge Yoshimoto et al. (2004)Rhodococcus equi Y50156 Metabolism of E1, E2, E3, EE2 Activated sludge Yoshimoto et al. (2004)Rhodococcus equi Y50157 Metabolism of E1, E2, E3, EE2 Activated sludge Yoshimoto et al. (2004)Rhodococcus equi ATCC 13557 Partial degradation of
EE2 in the presence of a cosubstrate ATCC O’Grady et al. (2009)
Rhodococcus erythropolis ATCC 4277 Partial degradation ofEE2 in the presence of a cosubstrate
ATCC O’Grady et al. (2009)
Rhodococcus rubber KC4 Conversion of E2 to E1 Activated sludge Yu et al. (2007)Rhodococcus sp. ED6 Metabolism of E2, E1 Soil samples from
agricultural fieldsKurisu et al. (2010)
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Sea. The strain H5 belongs to the genera Vibrio, and the strain S19- 1 belongs to the genera Buttiauxella.
5. Anaerobic biodegrada tion of estrogens
Early studies reported that several human intestinal bacteria and oral microorganism s are capable of degrading E2 under both aerobic and anaerobic conditions (Jarvenpaa et al., 1980; Kornman and Loesche, 1982; Ojanotko harri et al., 1991 ). Anaerobic degrada- tion of E2 to E1 was observed in activated sludge when amended with acetone (Lee and Liu, 2002 ). Using lake sediment as inocula,Czajka and Londry (2006) observed reversible conversion between E2 and E1 under anaerobic conditions. However, the further degra- dation of E1 was minimal. They also reported that E1 and E2 could be transformed into 17a-estradiol under anaerobic condition s, ex- cept the nitrate-reduci ng condition . Anaerobic biodegra dation ofEE2 by lake sediment microbial community was not observed (Czajka and Londry, 2006 ). However, anoxic biodegradat ion ofEE2 was observed in acclimated activated sludge with nitrate (Zeng et al., 2009a ). Two estrogen-degra ding denitrify ing bacteria,Denitratisom a oestradioli cum AcBE2-1T and Steroidobacter denitrifi-cans FST, were isolated from activated sludge and anoxic digested sludge, respectively (Fahrbach et al., 2006, 2008 ). Recently, Ivanovet al. (2010) reported that facultative anaerobic iron-reducing bac- terial isolates can remove estrogens: 27% for E1, 92% for E2, and 60% for E3. However, no degradation products were detected intheir study. Table 2 lists anoxic and anaerobic microorgan isms capable of transforming estrogens .
6. Microbial ecology of estrogen-degr ading consortia
Application of molecula r tools to characterize estrogen-degra d-ing microbial community structure was first reported by Yu et al.(2005). In their study, a quantitat ive fingerprinting method, called
real-time -t-RFLP, was used to characterize three 17a-estradiol-,E2-, and E1-degra ding microbial consortia, which were originally enriched from activated sludge. These enrichment cultures showed different community profiles. Interestingl y, seven specific ribo- types were present in all three microbial communi ties, suggesting that these ribotypes might be associated with estrogen degradat ion.
Microautorad iography–fluorescence in situ hybridiza tion was used to explore the diversity of E1-utilizing bacteria in activated sludge (Zang et al., 2008 ). Approximately 1% to 2% of the total cells,considered as E1-utilizing bacteria, were mostly affiliated with the long rod-shaped b-Proteobacteria and spherical and short rod- shaped c-Proteobact eria. By cloning and sequencing 16S rRNA genes of E2- and EE2-degr ading enrichme nt cultures, Mulleret al. (2010) obtained sequences belonging to Rhizobiales, Alcaligen-aceae, Sphingobacteriales, and Bacteroidete s. However, as their enrichme nts were supplemented with acetonitri le, the derived se- quences might be contributed from acetonitrile -degrading bacte- ria, not from estrogen- degrading bacteria.
Real-tim e PCR assays have been developed and applied toquantify specific estrogen- degrading bacteria in full-scale bioreac- tors. Using a real-time PCR assay specific to strain JEM-1, the average cell numbers of strain JEM-1 in a conventional activated sludge system and an oxidation ditch were 8.7 ± 4.4 and 40 ± 32 � 105 cells mg�1 mixed liquor volatile suspended solids (MLVSS), respectively (Hashimoto et al., 2010 ). Linear relation- ships between cell numbers of strain JEM-1 and the efficiencyof estrogen removal were observed. Recently , Roh and Chu (2010) used real-time PCR assays to investigate the presence oftwo estrogen- degrading bacteria, strains KC8 and ARI-1, in three activated sludge systems operating under different SRTs. Both strains were detected in all three systems; however, the average concentr ations of these strains were relative low: 0.2–1.9 � 105
cells mg�1 MLVSS for KC8 and 0.1 to 2.8 � 103 cells mg�1 MLVSSfor ARI-1.
Table 1 (continued)
Phylogenetic affiliation Degradation ability and mechanism Source of isolates References
Rhodococcus sp. ED7 Metabolism of E2, E1 Soil samples from agricultural fields
Kurisu et al. (2010)
Rhodococcus sp. ED10 Metabolism of E2, E1 Soil samples from agricultural fields
Kurisu et al. (2010)
Rhodococcus zopfii Y50158 Metabolism of E1, E2, E3, EE2 Activated sludge Yoshimoto et al. (2004)Rhodococcus zopfii ATCC 51349 Partial degradation of
EE2 in the presence of a cosubstrate ATCC O’Grady et al. (2009)
Bacteroidetes Flavobacterium sp. KC1 Conversion of E2 to E1 Activated sludge Yu et al. (2007)Flavobacterium sp. KC2 Conversion of E2 to E1 Activated sludge Yu et al. (2007)Sphingobacterium sp. JCR5 Metabolism of E1, E2, E3, EE2. Oral contraceptives
producing factory activated sludge
Ren et al. (2007)
Firmicutes Bacillus cereus Socransky 67 Conversion of E2 to unknown metabolites Dental plaque Ojanotkoharri et al.(1991)
Bacillus sp. E2Y1 Degradation of E1, E2 Activated sludge Jiang et al. (2010)Bacillus sp. E2Y2 Conversion of E2 to E1 Activated sludge Jiang et al. (2010)Bacillus sp. E2Y3 Conversion of E2 to E1 Activated sludge Jiang et al. (2010)Bacillus sp. E2Y4 Degradation of E1, E2 Activated sludge Jiang et al. (2010)Bacillus sp. E2Y5 Conversion of E2 to E1 Activated sludge Jiang et al. (2010)Staphylococcus aureus Conversion of E2 to E1 and vice versa ;
conversion of 16a-hydroxyestrone to E3Intestinal microorganisms Jarvenpaa et al. (1980)
Streptococcus faecalis Conversion of E2 to E1; conversion ofE1 to 16a-hydroxyestrone
Intestinal microorganisms Jarvenpaa et al. (1980)
Streptococcus mutans Ingbritt Conversion of E2 to E1 Dental plaque Ojanotkoharri et al.(1991)
Streptococcus mutans NCTC 10449 Conversion of E2 to E1 Dental plaque Ojanotkoharri et al.(1991)
Streptococcus sanguis NCTC 10904 Conversion of E2 to E1 Dental plaque Ojanotkoharri et al.(1991)
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7. Degradation and growth kinetics
Only a few studies have investigated estrogen degradat ion and growth kinetics by pure and mixed estrogen-de grading cultures (Table 3). Layton et al. (2000) used first-order degradation kinetics to describe the disappearan ce of testosterone, E2, and EE2 in acti- vated sludge (MLSS = 2165 mg L�1) within 24 h. The first-orderrate constants (k) for these compounds were determined based on the production rate of 14C-labeled CO2 from the 14C-labeled par- ent compounds . In another study, a zero-order kinetic was used todescribe the degradation of E2 observed in a mixed E2-degrading culture that was enriched with E2 and acetonitrile (Muller et al.,2010). The zero-order degradat ion rate constant was estimated to be 7.2 ± 0.2 mg E2 g�1 protein h�1. When additional carbon sources were present, a slower degradation rate constant of6.6 ± 0.4 mg E2 g�1 protein h�1 was observed. In addition, regard- less the presence of additional carbon source, much smaller degra- dation rate constants were observed for E1: 0.09 and 0.60 mg E1 g�1 protein h�1, with and without additional carbon sources, respectively.
Under aerobic conditions, initial transformation rates of E1 and E2 by an estrogen-degra ding bacterium Sphingom onas sp. CYH were determined (Ke et al., 2007 ). The determined transformation rates were linearly correlated to the initial concentratio ns of estro- gens. However , under anoxic conditions, the degradat ion of E1 bythe strain CYH followed Michaelis–Menten kinetics, with a maxi- mum degradation rate of 18.9 lg L�1 h�1 and a half-satu ration con- stant of 106 lg L�1. Monod-type degradation kinetics successfully described degradation of E1, E2, and testosterone by strain KC8 (Roh and Chu, 2010 ). The maximum specific substrate utilization rates are 0.37, 0.50, and 0.17 mg substrate mg�1 protein d�1 for
E2, E1, and testosterone, respectively . The half-saturat ion con- stants are 1.9, 2.7, and 2.4 mg L�1 for E2, E1, and testosterone,respectivel y. These kinetic paramete rs are comparable to those ofheterotrop hic microorgan isms in activated sludge.
Roh and Chu (2010) determined the doubling times and growth yields of strain KC8 when grown on different carbon sources. The doubling times of KC8 were 27, 61, 20, 24 and 29 h and the yields were 0.22, 0.24, 0.05, 0.04 and 0.42 g VSS g�1 BODL, when grown onE2, testosterone, glucose, sodium succinate and sodium acetate,respectivel y. A slightly shorter double time (18 h) was estimate dfor EE2-utilizing bacterium JCR5 when grown on EE2 (Ren et al.,2007). An E2-degrading, denitrify ing Denitratisom a oestradioli cum strain AcBE2-1 T has a doubling time of 23–25 h when grown onE2 (Fahrbach et al., 2006 ).
8. Estrogen degradation pathways
Studies on estrogen degradation pathways are limited and sporadic. In this section, based on published articles, we summa- rized two specific areas: Degradation pathways of E2 by aerobic bacteria and biotransformat ion of EE2 by bacteria, fungi, and algae.
8.1. Degradat ion pathways of E2 by aerobic bacteria
The first degradation pathway for E1 by a soil bacterium, Nocar-dia sp. E110, was proposed 40 years ago by Coombe et al. (1966). Intheir study, a dioxygenase was proposed to be responsible for the cleavage in the ring A of E1. This pathway has become as a part ofseveral aerobic E2 degradat ion pathways proposed in recent years.
Table 2List of anaerobic and anoxic steroid hormone-de grading and -transforming bacteria.
Phylogenetic affiliation Degradation ability and mechanism Source of isolates References
Alpha-proteobacteria Sphingomonas sp. CYH Degradation of E1, E2 under both aerobic and anoxic conditions
Artificial sandy aquifer
Ke et al. (2007)
Beta-proteobacteria Denitratisoma oestradiolicum AcBE2-1T Metabolism of E1, E2 under the denitrifying condition
Activated sludge Fahrbach et al.(2006)
Gamma-proteobacteria Steroidobacter denitrificans FST Metabolism of E1, E2 testosterone, 4-androstene-3,17-dione under the denitrifying condition
Anoxic digested sludge
Fahrbach et al.(2008)
Actinobacteria Actinomyces viscosus 378.5 Degradation of E2 and progesterone anaerobically
Subgingival plaque samples
Kornman and Loesche (1982)
Agromyces sp. LHJ3 Degradation of E2 and E3 with formation of E1 under aerobic condition; degradation of E2 with formation of E1 under anoxic condition
Artificial sandy aquifer
Ke et al. (2007)
Bacteroidetes Bacteroides fragilis Conversion of E1 to E2 and E1 to 16a-hydroxyestrone anaerobically
Intestinal microorganisms
Jarvenpaa et al.(1980)
Bacteroides gingivalis w Degradation of progesterone anaerobically
Subgingival plaque samples
Kornman and Loesche (1982)
Bacteroides gingivalis 167.5 Degradation of E2 and progesterone anaerobically
Subgingival plaque samples
Kornman and Loesche (1982)
Bacteroides gingivalis 208.1 Degradation of E2 and progesterone anaerobically
Subgingival plaque samples
Kornman and Loesche (1982)
Bacteroides melaninogenicus subsp. Intermedius 155.6 Degradation of E2 and progesterone anaerobically
Subgingival plaque samples
Kornman and Loesche (1982)
Bacteroides melaninogenicus subsp. Intermedius 166.5 Degradation of E2 and progesterone anaerobically
Subgingival plaque samples
Kornman and Loesche (1982)
Bacteroides melaninogenicus subsp. Intermedius 167.4 Degradation of E2 and progesterone anaerobically
Unclassified Iron-reducing bacteria with 16S rRNA gene 84%similar to Shewanella baltica
Degradation of E1, E2, E3 under iron- reducing condition
Anaerobic digester Ivanov et al. (2010)
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The first step of E2 degradation , as shown in Fig. 2, can be classifiedinto four specific groups: (i) hydroxylation of ring A at C-4, (ii)hydroxylation of saturated ring, (iii) dehydration of ring D at C-17, and (vi) dehydrogenati on of ring D at C-17.
(i) Hydroxylation of ring A at C-4 : Kurisu et al. (2010) detectedan intermediate metaboli te, 4-OH-E2, during E2 degradation by a soil estrogen- utilizing isolate, Sphingomona s sp. ED8 (hereafter referred as strain ED8). The detection of 4-OH- E2 suggested that E2 was hydroxylate d at C-4 of E2. In their study, they suggested that 4-OH-E2 was further degraded via meta cleavage.
(ii) Hydroxylation of saturated ring : Kurisu et al. (2010) also iden- tified a degradat ion pathway for E2 via ring hydroxylation on different positions, based on the detection of hydroxy- E2, keto-E2, keto-E1, and 3-(4-hydroxyphenyl)-2-hydroxy- prop-2-e noic acid. This pathway was observed in E2 degra- dation by the strain ED8. The formation of 3-(4-hydroxyphe nyl)-2-hydroxyprop-2-eno ic acid suggested that E2 degradation was not initiated from the A ring, but from asaturated ring (B, C, or D ring). However, how the saturated ring was cleaved was not further investiga ted in their study.
(iii) Dehydrat ion of ring D at C-17 : By using Nitrosomonas euro- paea, Nakai et al. (2011) observed a new intermediate metabolite – estra-1,3 ,5(10),16-tetraen-3-ol (estratetraenol,E0). E0 was formed by dehydration of ring D at C-17 position of E2. Although E0 still possesse s estrogeni city, N. europaea could further degrade E0 into non-estro genic compounds (Nakai et al., 2011 ).
(iv) Dehydrogena tion of ring D at C-17 : Studies have demon- strated the formation of E1 from E2 biodegradat ion (Yuet al., 2007 ). E1 was formed through a dehydrogenati on pro- cess on the D ring C-17 position of E2. Degradation pathway of E1 in a soil isolate, Nocardia sp. E110, was first proposed by Coombe et al. (1966). Through hydroxylation at the A ring C-4 position of E1, E1 can be transformed into 4-OH E1,which can be further degraded via meta-cleavage (Kurisu
et al., 2010 ). Another degradation pathway was suggested by Lee and Liu (2002). By studying E2 degradation in mixed sewage bacteria, Lee and Liu (2002) detected a new metab- olite, X1, containing a lactone at ring D. X1 is believed tobe further degraded by entering the TCA (tricarboxylic acid)cycle.
8.2. Biotransf ormation of EE2 by bacteria, fungi, and algae
Biodegra dation of EE2 by fungal species has been demonstrat ed(Shi et al., 2002; Blánquez and Guieyssea, 2008 ). Shi et al. (2002) re-ported an unidentified polar metabolite from EE degradat ion by anEE2-degr ading fungus, Fusarium proliferratum strain HNS-1. The strain was isolated from a cowshed sample. The potential of different fungal enzymes to catalyze oxidation of natural and synthetic estro- gen has also been investigated. For example, manganese peroxidase (Suzuki et al., 2003 ), horseradi sh peroxidase (Auriol et al., 2006,2008), and laccase (Tanaka et al., 2001; Suzuki et al., 2003; Auriol et al., 2007, 2008 ) from fungi, such as Phanerochaete chrysosporiu m,Trametes sp., and Pycnopor us coccineus , have shown to degrade EE2. However, transformat ion products of EE2 were not successfully determined in these studies, leading to a knowledge gap in the deg- radation pathways for EE2 by fungus species.
Some bacteria and microalgaes were able to transform EE2 (Renet al., 2007; Della Greca et al., 2008 ). The first step of EE2 degradation can be classified into several different groups: (a) A ring C-2 hydrox- ylation, (b) A ring 3-OH conversion to 3-keto, (c) B ring C-6 hydrox- ylation, (d) D ring C-17 conversion to keto, and (e) conjugat ion ofEE2. Degradation pathways for EE2 are summari zed in Fig. 3.
(a) A ring C-2 hydroxylatio n: Degradation of EE2 was observed inan enriched nitrifying culture (Yi and Harper, 2007 ). In this study, 2-OH EE2 was detected, suggesting the hydroxylation of A ring at C-2 position of EE2. Cleavage of ring A proceeded before other rings (B, C, or D rings). Results of their study suggested that EE2 was cometabolical ly transformed bynitrifying bacteria.
Table 3Estrogen degradation kinetic parameter s and cell doubling time .
Mixed or pure cultures Kinetic parameters Values References
– 0.0152 ± 0.0021 min �1 0.0002 ± 0.0000 min �1 Laytonet al.(2000)
E2-degrading consortium enriched with E2 and acetonitrile
Specific degradation rate (zero-orderrate constant)
0.09 and 0.60 mg E1 g�1 protein h�1, with and without additional carbon sources (acetone and acetonitrile), respectively
7.2 ± 0.2 and 6.6 ± 0.4 mg E2 g�1
protein h�1, with and without additional carbon sources (acetoneand acetonitrile), respectively
– Mulleret al.(2010)
Sphingomonas sp. CYH Michaelis–Mentenkinetics under anoxic conditions
Maximum specific degradation rate 18.9 lg L�1 h�1; half- saturation constant 106 lg L�1
– – Ke et al.(2007)
Sphingomonas strain KC8 Monod kinetics and cell doubling time
Maximum specific substrate utilization rates 0.5 mg E1 mg�1
protein day �1; half-saturation constants 2.7 mg L�1
Maximum specific substrate utilization rates 0.37 mg E2 mg�1
protein day �1; half-saturation constants 1.9 mg L�1; doubling time 27 h
– Roh and Chu(2010)
Sphingobacterium sp. JCR5 Cell doubling time – – 18 ha Ren et al.(2007)
Denitratisoma oestradiolicum strain AcBE2-1T
Cell doubling time – 23–25 h – Fahrbachet al.(2006)
a Estimated from the data in the article (Ren et al., 2007 ).
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(b) A ring 3-OH conversion to 3-keto : This degradation pathway was observed in microalgae. Della Greca et al. (2008) founddifferent microalgae species, including Selenastrum capricor- nutum, Scenedesmu s quadricau da, Scenedesmus vacuolatus ,and Ankistrodesmus braunii , were able to transform EE2 into different transformat ion products through hydroxylation and glucosylation. S. quadricau da could convert 3-OH into 3-keto in A ring of EE2.
(c) B ring C-6 hydroxylatio n: This EE2 degradation pathway was only reported in microalgae (Della Greca et al., 2008 ). Ankis-trodesmu s braunii was reported in the study to cause B ring C-6 hydroxylation.
(d) D ring C-17 conversion to keto : Ren et al. (2007) identifiedseveral transformat ion metabolites during EE2 biodegrada- tion by an EE2-utilizi ng isolate, strain JCR5. These metabo- lites, including 3,4-dihyd roxy-9,10-seco androsta-1,3,5(10)-triene-9.17- dione, 2-hydrox y-2,4-dienevaler ic acid, and 2-hydroxy-2,4 -diene-1,6-d ioic acid, were detected by mass spectrum analysis. The initial step of EE2 degradation bythe strain JCR5 is to oxidize the C-17 of EE2 into a ketone,E1. As shown in Fig. 3, the C-9 of E1 is subsequently hydrox- ylated and ketonizated, resulting in cleavage of the B ring ofE1. Followed by hydroxylation in the A ring of the metabo- lite, 3,4-dihydro xy-9,10-sec oandrosta-1,3,5(10)-triene-9.17-dione was formed. Several subsequent degradation steps, including meta-cleavage of A ring via dioxygenase,lead to production of end products, CO2 and water.
(e) Conjugation of EE2 : This pathway was observed in microal- gae, Scenedesmus capricorn utum (Della Greca et al., 2008 ).The glucosylation of EE2 was detected on the C-3 position.S. capricornutum also showed the ability to cause hydroxyl- ation on the C-2 and C-6 positions of EE2.
9. Future research
Given the active research on isolation and characteri zation ofsteroidal hormone- degrading bacteria from wastewa ter, knowl- edge on microbiology of estrogen degradation has accumulate d inrecent years. However, we have just scratched the surface of the body of this knowledge. For example, steroidal hormones are not only detected in wastewater but also in a variety of environm ents,including seas and manure (Atkinson et al., 2003; Combalbert and Hernandez- Raquet, 2010 ). Current knowledge on the biodegrada- tion potential of sorbed estrogens in biosolids used for land applica- tions and/or landfills is limited. As most of reported estrogen degraders were isolated from wastewater , more research is needed to isolate steroidal hormone-de grading bacteria from these differ- ent environm ents. In addition, information on anaerobic/anox icbiodegra dation of estrogens, in terms of degradative strains and metaboli tes and pathways , is limited. Considering that 99% of bac- teria are unculturable in laborator y settings, little is known about the diversity and the degradat ive capacity of these bacteria in acti- vated sludge and other environments. Stable isotope probing (SIP)is a powerful tool used to study active contaminan t-degrading
Fig. 2. Degradation pathways of E2 by aerobic bacteria.
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microbial community in complex environments , like groundwater and soil microcosms (Yu and Chu, 2005; Roh et al., 2009 ). Using commerciall y available 13C-labeled E2 (C-3 and C-4 positions), se- ven unique bacterial clones were obtained from an E2-degrading consortium (Chu, 2010 ). These clones are completely different from known estrogen-degra ding isolates. Studies using SIP with fully 13C-labeled estrogens will allow researchers to probe the microbiol- ogy of estrogen degraders in different environments .
Future research efforts on elucidating the genes involved inestrogen biodegradation and understand ing the regulation and expression of these catabolic genes are needed. While the catabolic genes encoding enzymes responsib le for testoster one biodegrada- tion have been studied extensively, little is known about the genes associated with estrogen biodegra dation. Given the advances insequencing and omic techniques, whole genome sequencing and omic analysis of degradative strains is possible. The genome infor- mation can offer a comprehensive understand ing of the function and growth of the degradative bacteria of interest. For example,the whole genome sequencing of E2-utilizing bacterium strain KC8 (Hu et al., 2011 ) and strain SJTE-1 (Liang et al., 2012 ) were re- cently completed. The sequence suggests that these two strains have a very diverse catabolic ability and a unique ability to adapt and/or survive in different environm ents.
The 16S rRNA genes of estrogen-degra ding bacteria have been successfully used as biomarkers to study the presence and abun- dance of known estrogen-degra ding bacteria in different lab-scale bioreactors and WWTPs. However , such an approach is limited bythe availabili ty of 16S rRNA gene sequence s from isolates and from uncultivated estrogen degraders, leading to underestimation ofestrogen-de grading bacteria in a given environment. Using key functional genes responsible for estrogen biodegra dation would be a better approach to overcome this limitation as well as to
explore the diversity of estrogen-degra ding microbial community.For the case of strain KC8, Rieske dioxygenase and catechol 2,3- dioxygenase have been implicated for their roles in estrogen bio- degradat ion. Further studies are needed to confirm their degrada- tive functions during estrogen degradat ion. Using proteom icanalysis, a recent study revealed that Stenotropho monas maltophilia strain ZL1 might convert E1 to tyrosine and utilize the tyrosine for its protein biosynthesis (Li et al., 2012 ). This study suggests that omic techniques are useful tools for future estrogen biodegrada- tion research.
Enhancin g estrogen removal during WWTP processes is a new challenge for environmental engineers. The number of several known estrogen-degra ding bacteria is low in activated sludge sys- tems operating under different SRTs (Hashimoto et al., 2010; Roh and Chu, 2010 ). Despite removals of E2 in the bioaugment ed biore- actors were high (>99%), bioaugmentati on of estrogen-de grading bacteria in a full-scale activated sludge system might be challeng- ing, because the number of these aerobic degraders are subjected to gradually decrease with time (Roh and Chu, 2011 ). Therefore,searching for effective methods, possibly using membrane bioreac- tors and/or immobilization techniqu es, would be a key for sustain- able estrogen removal through bioaugment ing known estrogen degraders .
Biodegra dation kinetics of steroidal hormones and their degra- dation metabolites is essential information for assessing transfor- mation and transport of estrogens in the environment. Due tothe low ambient concentrations of estrogens in the environment,further studies are necessary to confirm if the reported degradation kinetics are relevant to and suitable for predicting the behavior ofhormones in the environment. The estrogeni c potential of estrogen metaboli tes is an important aspect of estrogen degradation, either biotically or abiotically. Although these transformed byproducts
((
((
( (
(((
(
Fig. 3. Degradation pathways of EE2 by bacteria and algae.
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are different from their parent compounds , they might still possess estrogenicit y. Previous studies have shown that fungal and algae can transform estrogens into different metabolites , and nitrite can react with estrogens to form nitro-estrogen s. However, it isnot clear if those metabolites still have estrogeni c potential and/ or can be further biodegraded in the environment. Therefore, stud- ies are needed to investigate the fate of these transformed byproducts.
Over a decade, significant progress has been made in elucidat- ing the mechanis ms of estrogen biodegradat ion. Yet, we have just scratched the surface of the body of the knowledge on estrogen biodegradat ion. Further research is still required to enhance our understand ing on the microbiology of estrogen biodegradation –a needed knowledge base to assist the development of effective re- moval processes for estrogens.
Acknowled gements
This work is supported by the Special Program for Key Basic Re- search of the Ministry of Science and Technology , China (2010CB434802); the Science and Technology Planning Project ofXiamen, China (3502Z20102017, 3502Z20111 049); and the CAS/ SAFEA Internationa l Partnership Program for Creative Research Teams (KZCX2-YW-T08).
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