Appl Microbiol Biotechnol DOI 10.1007/s00253-014-5826-0

MINI-REVIEW

Cometabolic degradation of organic wastewater micropollutants by activated sludge and sludge-inherent microorganisms Klaus Fischer & Marius Majewsky

Received: 28 February 2014 / Revised: 10 May 2014 / Accepted: 13 May 2014 # Springer-Verlag Berlin Heidelberg 2014

Abstract Municipal wastewaters contain a multitude of organic trace pollutants. Often, their biodegradability by activated sludge microorganisms is decisive for their elimination during wastewater treatment. Since the amounts of micropollutants seem too low to serve as growth substrate, cometabolism is supposed to be the dominating biodegradation process. Nevertheless, as many biodegradation studies were performed without the intention to discriminate between metabolic and cometabolic processes, the specific contribution of the latter to substance transformations is often not clarified. This minireview summarizes current knowledge about the cometabolic degradation of organic trace pollutants by activated sludge and sludge-inherent microorganisms. Due to their relevance for communal wastewater contamination, the focus is laid on pharmaceuticals, personal care products, antibiotics, estrogens, and nonylphenols. Wherever possible, reference is made to the molecular process level, i.e., cometabolic pathways, involved enzymes, and formed transformation products. Particular cometabolic capabilities of different activated sludge consortia and various microbial species are highlighted. Process conditions favoring cometabolic activities are emphasized. Finally, knowledge gaps are identified, and research perspectives are outlined.

K. Fischer (*) Department of Analytical and Ecological Chemistry, University of Trier, Behringstr. 21, 54296 Trier, Germany e-mail: [email protected] M. Majewsky Engler-Bunte-Institut, Chair of Water Chemistry and Water Technology, Karlsruhe Institute of Technology (KIT), Engler-Bunte-Ring 1, 76131 Karlsruhe, Germany e-mail: [email protected]

Keywords Wastewater treatment . Biodegradation . Activated sludge . Cometabolism . Enzymes . Emerging pollutants . Pharmaceuticals and personal care products . Antibiotics . Estrogens . Nonylphenol

Introduction Nowadays, municipal wastewaters contain a multitude of organic micropollutants including emerging pollutants, e.g., pharmaceuticals, personal care products, antibiotics, flame retardants, etc., as well as pollutants with a long emission history, e.g., tensides, phthalates, herbicides, and pesticides (Richardson and Ternes 2011; Fischer et al. 2012). These contaminants might pose risks for aquatic ecosystems and for human health, if they are not completely eliminated during wastewater treatment. Historically, wastewater treatment plants (WWTPs) are designed to remove the bulk of the organic carbon load, nitrogen, and phosphorus, not to tackle xenobiotics. Except for plants working with an additional advanced treatment step (activated carbon, photolytic, or oxidative treatment), the removal of trace pollutants is rather a side effect of a proper plant operation. For many micropollutants, especially polar ones, physicochemical elimination processes are of minor importance and their overall removal is largely governed by the efficiency of microbial degradation performed in activated sludge (AS) systems (Reemtsma and Jekel 2006). Microbial substrate transformation encompasses the metabolic type, maintaining the vital functions of the degraders including their growth and reproduction, and the cometabolic type. The latter is defined as “the transformation of a non-growth substrate in the obligate presence of a growth substrate or another transformable compound” (Dalton and Stirling 1982). Occasionally, the term “fortuitous degradation” is in use.

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For several reasons, a distinct differentiation between both processes is not always possible for mixed microbial communities such as activated sludge consortia, and various transitional forms have to be taken into account: &

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Several xenobiotics, e.g., chlorinated compounds, are degradable via both metabolic and cometabolic pathways, depending on the species composition of the microbial community and on the reaction conditions (Çeçen et al. 2009). Cometabolic and metabolic reaction steps might be closely interrelated and substitutable, since they are part of a metabolic network, evolved as a specific metabolic competence of the whole microbial community. The adaptation of microbes to new substrates proceeds via the development of new metabolic tools, e.g., the evolution of structurally altered enzymes, allowing for the replacement of a cometabolic process by a metabolic one. Recent investigations locate promiscuous enzymes at the interface between intended (metabolic) and unintended (cometabolic) degradation reactions (Hult and Berglund 2007; Elias et al. 2008; Babtie et al. 2010).

Nevertheless, as far as possible, a discrimination between metabolic and cometabolic processes is meaningful and required in many cases, since it has implications for the rationalization of pollutant removal efficiencies and for the optimization of operation conditions of WWTPs. For instance, cometabolic processes might specifically depend on the nature and amount of the bioavailable substrate (growth substrate), i.e., organic wastewater macrocomponents (Liu et al. 2013b) as well as on the concentration of inorganic solutes, i.e., ammonium. As a consequence, the fate of micropollutants is inherently linked to that of macropollutants (DelgadilloMirquez et al. 2011). Cometabolically generated intermediates and metabolites might differ from those metabolically formed possibly resulting in “dead-end” metabolites (Nalli et al. 2006; Evangelista et al. 2010). In certain cases, products of cometabolic processes exhibit higher toxicity than metabolically formed ones (Evangelista et al. 2010; Gabriel et al. 2012). As far as cometabolic activities are provided by specific community subgroups (heterotrophs and autotrophs, e.g., ammonia oxidizing bacteria), the plant process parameter “sludge retention time (SRT)” influences their relative sludge proportions and thus the biomass-related pseudo first-order degradation rates of the targeted micropollutants (Helbling et al. 2012; Xia et al. 2012). Finally, since “cometabolism results from the lack of specificity of enzymes and cofactors” (Criddle 1993), some cometabolic degradation processes might be performed by enzymes and cofactors other than those responsible for metabolic transformations of the same substrate if such a pathway exists. In this context, factors regulating the diversity, enrichment, stability, accessibility, conformational flexibility,

and activity of extracellular enzymes located at or in the sludge floc matrix might be of pivotal importance. It is obvious that attempts to explain the high variability of activated sludgerelated biodegradation rates of various micropollutants will fall short if those aspects are not taken into account (Helbling et al. 2012; Pomiès et al. 2013). Since biodegradation pathways are characterized by a unique combination of tackled substrates, active enzymes, and formed metabolites, the analytical identification of transformation products is an indispensable prerequisite for an in-depth elucidation of the biotransformation processes. Without this knowledge, metabolic competences of different microbial consortia and the related kinetic parameters might not be comparable. Nevertheless, many studies dealing with the fate of xenobiotics during sewage treatment report only disappearance of the original substances, partially measured by sum parameters exclusively. The total concentration of natural organic compounds in raw sewage surpasses that of the xenobiotics by several orders of magnitude. Thus, it is speculated that the biodegradation of the latter is mainly of cometabolic nature (Ternes and Joss 2006; Harper et al. 2008). If the natural substrate and the xenobiotic compound compete for the same active center of an enzyme, high concentrations of the natural substrate may lead to a reduced transformation of the micropollutant due to competitive enzyme inhibition (Alvarez-Cohen and Speitel 2001; Joss et al. 2004; Plosz et al. 2010). Further, the cometabolism of structurally non-analogous substrates, i.e., cometabolism without competitive inhibition, is also recognized (Brandt et al. 2003). Since the investigation of cometabolic processes requires control over amount and composition of the substrates as well as the application of isotopically labeled substances (if mass balances are an aim), very few attempts were made on the WWTP full scale level, e.g., by Lozano et al. (2013). Most experiments were conducted at laboratory or pilot scale level using either genuine-mixed microbial cultures from full-scale WWTP or manipulated cultures, e.g., enriched or fortified with potentially “high performing” specialized microbial species. The cometabolic degradation of halogenated aliphatic and aromatic compounds including chlorinated solvents and chlorophenols is comprehensively examined in the context of groundwater and soil remediation and summarized in several review articles (Chang and Alvarez-Cohen 1995; Daniel et al. 2001; Çeçen et al. 2009; Sayavedra-Soto et al. 2010; Wendlandt et al. 2010). Thus, these compounds are not considered in this review as well as azo dyes which play a considerable role as pollutants of some industrial but usually not of domestic wastewaters (Ambrosio and Campos-Takaki 2004; Sugumar and Sadanandan 2010; Pajot et al. 2011). Reviewing the cometabolic transformation of wastewater micropollutants, the focus in this study is mainly on pharmaceuticals and personal care products (PPCP), antibiotics,

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hormones, and tensides. Further, the identification of process conditions favoring the cometabolic degradation of distinct groups of pollutants is one of the main topics of this literature survey. Another central aspect is the characterization of microbial species and of related enzymes involved in contaminant transformation. Lastly, knowledge gaps, new scientific approaches, and research perspectives are briefly outlined.

Degradation of various pollutant groups and of individual compounds Antibiotics and pharmaceutically active compounds (PhACs) Antibiotics Among the variety of pharmaceuticals usually detected in wastewaters, sulfamethoxazole (SMX) is often taken as a lead substance for sulfonamide antibiotics and therefore frequently integrated in micropollutant degradation studies. While its overall removal efficiencies in WWTPs have been the subject to many investigations, only a few biodegradation studies have focused on the aspect of cometabolism mediated by isolated microorganisms or activated sludge consortia (ASC). Readily biodegradation of SMX was observed in flask reactors at lab-scale performed in a semi-continuous mode with activated sludge after a lag phase of 14 days at a SMX concentration of 10 mg L−1 (Müller et al. 2013). The addition of readily biodegradable substrate resulted in enhanced SMX removal under aerobic conditions. Acetate and NH4NO3 were continuously provided as primary substrates but were also readily consumed and may have led to diauxic degradation rather than cometabolism, i.e., that SMX is only used as nutrient source as soon as the readily available primary substrates have been oxidized. Two possible degradation pathways were proposed (Fig. 1): 1. The formation of 3-amino-5-methyl-isoxazole and 4hydroxy-sulfamethoxazole, in which the p-amino group has been substituted by a hydroxyl group, when SMX is the only source of carbon and nitrogen; 2. The additional formation of N-acetyl-sulfamethoxazole and sulfanilic acid, when SMX is supplied with a growth substrate. The formation of the latter together with 3amino-5-methyl-isoxazole can be seen as indication for hydrolytic cleavage of the sulfonamide group. Nonetheless, the ambiguous picture and the question whether primary substrate enhances or competes SMX breakdown as well as the linkage to (non-specific) enzymes and the capabilities of single microbial species still require further research and clarification at this point. Obviously, bacteria capable of degrading sulfonamide antibiotics are usually

present in activated sludge consortia and can be enriched by providing SMX as the only or dominant substrate. Certainly, the bacteriostatic nature of SMX can cause a shift of the microbial composition (Collado et al. 2013), which however must be distinguished from (co-)metabolic adaption processes. Hence, species-specific investigations may provide further insight into the functioning of activated sludge communities. In this context, Larcher and Yargeau (2011) tested seven species (Bacillus, Pseudomonas, Rhodococcus) as pure cultures known to exist in activated sludge communities for SMX degradation in the presence and absence of glucose. While the cell density of all species increased due to growth on the monosaccharide, an increasing SMX removal could only be observed for three species. Further, metabolite formation was observed for Rhodococcus equi but only in the presence of glucose assuming the production of enzymes specific for this species such as arylamine N-acetyltransferases (specificity for aromatic amines), amidases, urethanase (hydrolizes anilides), or N-acetyl-phenylethylamine hydrolase (hydrolizes N-acetylated compounds), which were proposed to lead to N-acetylated SMX and N-hydroxy-acetylated SMX as transformation products. Among five isolated strains (phyla Actinobacteria and Proteobacteria), Rhodococcus sp. was also reported to mineralize the 14C-labeled aniline moiety of SMX, when supplying the latter as the sole source of carbon (Bouju et al. 2012). Prior to this, the sludge was adapted to a synthetic medium spiked with SMX (100 μg L−1) operated in a membrane bioreactor (MBR) for 10 months. Pharmaceutically active compounds (PHACs) Similar phenomena were reported for PhACs. Nitrifiers taken from a municipal WWTP and enriched using a lab-scale reactor operated with minimum salt medium (MSM) and ammonium for 2 months showed enhanced biodegradation of numerous PhACs as compared to conventional activated sludge (CAS). The biotransformation increased with increasing addition of ammonium and acetate as primary substrates (Tran et al. 2009). This enhanced removal is explained by a combination of a widened microbial diversity and a broader spectrum of non-specific enzymes such as monooxygenase and dioxygenases. Another study reported faster removal for the cholesterol-lowering drug statin when additional organic material was present indicating a cometabolic degradation mechanism (Ottmar et al. 2012). On the contrary, findings by Urase and Kikuta (2005) indicated better removal of 15 common PhACs during low organic substrate conditions in batch experiments with CAS. The sludge was not acclimatized to PhACs before, and the different findings might therefore be explained by the different microbial consortium and enzyme induction. Generally, with regard to the acclimatization of microbial communities to PhACs, it remains unclear whether the concentrations occurring in WWTPs—usually in the nanogram to

Appl Microbiol Biotechnol Fig. 1 Selected transformation products of sulfamethoxazole proposed to be formed via different transformation pathways A, B, C, and D: cometabolic (Gauthier et al. 2010; Larcher and Yargeau 2011; Müller et al. 2013); B and C: metabolic (Müller et al. 2013)

microgram per liter range—are high enough to promote such population shifts, or whether the dominating biodegradable organic load is the (only) determining factor. The latter case would support that non-specific enzymes greatly govern PhAC degradation due to their structural similarity to the enzymes’ natural target substrate. In this regard, non-specific enzymes and general heterotrophic biomass activity were also hypothesized to govern cometabolic breakdown of four PhACs different in structure (Majewsky et al. 2011). Significant faster degradation kinetics could be observed at a high heterotrophic activity (SRT 6 days) as compared to a sludge from a different WWTP with a low heterotrophic activity (SRT 54 days). Experiments were carried out under identical conditions with synthetic substrate but without any prior adaption or cultivation. Roh et al. (2009) used pure ammonia-oxidizing bacteria (AOB) cultures of Nitrosomonas europaea and observed biodegradation for triclosan and bisphenol A but not for ibuprofen. Experiments with and without allythiourea (ATU), an inhibitor for the nonspecific enzyme ammonia monooxygenase (AMO), suggested that AMO might be responsible for triclosan and bisphenol A degradation via cometabolic reactions. The low substrate specificity of AMO usually facilitates oxidative

cometabolic breakdown of a variety of compounds such as aliphatic and aromatic hydrocarbons. Additionally, experiments with two nitrifying activated sludges (NAS) were carried out, in which degradation was found for all three compounds in the presence and absence of ATU within a few days. It was concluded that heterotrophs and non-ammoniaoxidizing microorganisms in mixed activated sludge communities can also be involved in PhAC breakdown reactions. Cometabolic PhAC transformation was suggested to be governed by autotrophs via AMO, while heterotrophs can act upon PhACs via both metabolic and cometabolic pathways (Tran et al. 2013). Sathyamoorthy et al. (2013) investigated the cometabolic biodegradation of the betablocker atenolol (15 μg L−1) during nitrification in batch experiments. Biodegradation rate constants were found to be approximately four times greater in batches without nitrification inhibition by ATU. The effectivity of ATU was ensured by nitrite and nitrate controls. From experimental and modeling results, it was concluded that ammonia-oxidizing bacteria contribute about 7–17 % to the biodegradation rate constant of atenolol. In contrast, Helbling et al. (2012) statistically evaluated the possible correlation of the oxidative transformation rates of ten micropollutants with WWTP process parameters and

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acetylene-inhibited monooxygenase activities including AMO. The authors found a correlation with ammonium removal but not with AMO activity. At the consortium level, Kraigher et al. (2008) investigated the effect of the five acidic pharmaceuticals ibuprofen, naproxen, ketoprofen, diclofenac, and clofibric acid on the microbial AS composition taken from a municipal WWTP by 16S rRNA gene analysis. Therefore, the authors used labscale bioreactors continuously fed with artificial wastewater and PhAC concentrations at 5 to 50 μg L−1 for each compound for a period of 2 years, accompanied by a control without PhACs. Subsequently, concentrations were increased up to 500 μg L−1 in other bioreactors for 2 months using the adapted sludge. It was concluded that mixed pharmaceutical concentrations of 50 μg L−1 caused shifts in the bacterial community structure and also reduced bacterial diversity. The authors state that the (co-)metabolic degradation of these compounds, mostly containing aromatic structures, “is most likely the inherent ability of the sludge communities”. Another study compared the biodegradation of five common acidic PhACs using the standard biodegradation tests (ISO 7827, municipal WWTP inoculum) with and without external carbon source (milk powder) at concentrations in the lower milligram per liter range (Quintana et al. 2005). Ketoprofen was found to be degraded in the absence of the additional carbon source after a lag phase of 10 days but remained stable when milk powder was added. In contrast, cometabolic degradation was observed for naproxen, ibuprofen, and bezafibrate and additionally confirmed by the emergence of the corresponding transformation products. For ketoprofen, a degradation pathway known for biphenyls via dioxygenases was proposed. The hydrolytic cleavage of the amide bond for bezafibrate gave indication about the involvement of (unspecific) amidases, as also suggested by Helbling et al. (2010b).

Natural and synthetic estrogens Natural estrogens, e.g., estrone (E1), 17β-estradiol (E2), and α-estriol (E3), as well as synthetic, e.g., 17α-ethinylestradiol (EE2), estrogens are assessed as endocrine-disrupting compounds (EDCs) with potential adverse effects on aquatic organisms. They are considered to be major contributors to the estrogenic activity of some WWTP effluents (Gutendorf and Westendorf 2001; Pauwels et al. 2008). Thus, factors and conditions affecting their elimination during wastewater treatment are intensively investigated (Clouzot et al. 2008; Cajthawl et al. 2009; Limpiyakorn et al. 2011; Yu et al. 2013). Due to the close relation between estrogen biodegradation and microbial growth, the applicability of Monod-type degradation kinetics (Estrada-Arriaga and Mijaylova 2010)

and due to the strong dependence of the degradation on the amount and composition of easily available substrates, the hypothesis of a predominance of cometabolic degradation processes, at least in EE2 elimination, is widely accepted. Generally, biodegradation of estrogens requires aerobic conditions (Joss et al. 2004). Elimination under anaerobic conditions is restricted to sludge adsorption (Mes et al. 2008), which was also ascertained under aerobic conditions after chemical inactivation of the sludge (Clara et al. 2004; Xu et al. 2008). Highest degradation rates were achieved with NAS (Andersen et al 2003; Limpiyakorn et al. 2011). Under these conditions, close relations between the degradation rate, the concentration of biomass (Cao et al. 2008), and the nitrification rate (Yi and Harper 2007; Dytczak et al. 2008) were established. An increase of the organic wastewater load depressed estrogen removal (Ren et al. 2007), being effective at restricted ammonium supply only. At an ammonium concentration of 30 mM, the EE2 transformation (initial concentration 3.5 mg L−1) was not impeded by the highest chemical oxygen demand (COD) concentration applied (140 mg L−1) (Likitmongkonsakun and Limpiyakorn 2013). These phenomena together with the strong decline of EE2 elimination in the absence of ammonium and after inhibition of the AMO by addition of ATU led to the conclusion that AOB play a pivotal role in EE2 degradation (Pholchan et al. 2008; Forrez et al. 2009; Maeng et al. 2013). This reasoning is further supported by the outcome of the operation of a MBR filled with a nitrifier enrichment culture (NEC) (De Gusseme et al. 2009). Treating a synthetic influent with an EE2 concentration of 83 ng L−1 resulted in a removal efficiency of 99 % at an optimal NH4+-N concentration of 1.0 mg L−1. Parallel tests with the NEC in batch reactors revealed a strong decline of EE2 degradation after ATU addition. The enrichment of heterotrophic bacteria did not promote EE2 removal. In this case, EE2 elimination could be attributed to biomass adsorption. Nevertheless, due to several opposite findings, this reasoning is not without controversy (Gaulke et al. 2008; Zhou and Oleszkiewicz 2010). Yi and Harper (2007) demonstrated by means of an AMO containing protein extract and by balancing the NADH demand related to the enzymatic EE2 degradation that AMO and not a dioxygenase, stemming from heterotrophic bacteria, was responsible for the estrogen transformation in an enriched culture of autotrophic ammonia oxidizers. They detected OH-EE2, sulfate-EE2, and a compound with cleaved ring A as EE2 transformation products. They also formulated the idea that autotrophic and heterotrophic bacteria might be able to closely cooperate in estrogen removal. An indication for that is the absence of intermediates in mixed cultures which otherwise would accumulate in pure AOB cultures (Shi et al. 2004). The operation of two sequencing batch reactors (SBRs) with different activated sludge communities elucidated a variable contribution of nitrifiers to the overall E2 and EE2 degradation. The EE2 elimination dropped in the SBR1 from

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87.7 to 76.9 % after ATU addition and in the second reactor from 33.7 to 15.9 %, respectively. Despite higher proportions of AOB to the biomass of the peptone-fed SBR1 with the second one being glucose fed, most of E2 and EE2 were transformed by heterotrophic bacteria (Racz et al. 2012). Recently, Khunjar et al. (2011) shedded light in the synergistic degradation performance of both bacteria groups. They assayed the EE2 degradation by different pure, enriched, and NAS cultures in flow-through reactors under various feed conditions at realistic EE2 concentrations (ng L−1). The dioxygenase activity in heterotrophic cultures was downregulated in an experimental subset by feeding with acetate only and upregulated (enzymatic activity 23–63 times higher than in the downregulated subset) by additional feeding with benzoate and toluene. The formed degradation products differed depending on the dioxygenase activity in heterotrophic cultures and between the latter and the autotrophic cultures. The basal expression of heterotrophic dioxygenase enzymes was sufficient to achieve a high degree of EE2 transformation. 2Nitro-EE2 and 4-nitro-EE2 were exclusively formed in AOB cultures presumably via an abiotic mechanism but degraded by heterotrophs to a high extent. Further, AOBs degraded EE2 five times faster than heterotrophs. After inactivation of AOBs, the heterotrophic microorganisms adapted within a few days filling the gap in the EE2 degradation process chain.

Nonylphenols The biodegradation of nonylphenol ethoxylates (NPEOs), an important group of non-ionic surfactants, during sewage treatment leads to the formation of nonylphenols (NPs) and related nonylphenolic compounds, e.g., nonylphenoxyacetic acid. This process reconstitutes the mixture of various NP isomers which are ingredients of the commodity chemical NP, applied for the industrial synthesis of NPEO. As a result of its synthesis process, NP is a mixture of more than 150 isomers differing in the structure and in the ring-binding position of the alkyl (mainly nonyl) moiety (Ieda et al. 2005; Eganhouse et al. 2009). Depending on the producer, 4-nonylphenols account for approximately 86–94 % of the technical mixture, with 2-nonylphenols and decylphenols making up 2–9 and 2– 5 %, respectively. Thus, the pattern of nonylphenolic compounds arising during the wastewater treatment process and in the aquatic environment is shaped by the isomer distributions of the NP surfactants and by the environmental properties including biological and abiotic degradability of the specific isomers. Consequently, in-depth studying the fate of nonylphenolic compounds during sewage treatment requires an isomer-selective approach. The biodegradability of the NPs is lower than that of the NPEOs not least because of the stronger sorption of the former. Nevertheless, several heterotrophic bacteria, Sphingonomas sp. TINP3, Sphingonomas

cloacae, and Sphingonomas xenophagum strain Bayram were isolated from activated sludge for their ability to grow with four different NPs as sole carbon and energy source (Corvini et al. 2006; Giger et al. 2009). Studying the isomeric specific NP degradation by S. xenophagum Bayram, it was ascertained that isomers with hydrogen atoms at the benzylic position were differently transformed from those with a quaternary carbon atom at that position (Gabriel et al. 2005a, b, 2012; Kohler et al. 2008). In the case of the latter which served as growth substrates, the nonyl group is detached from the phenolic unit and transformed into a nonyl alcohol, whereas the aromatic unit is oxidized to yield hydroquinone (pathway A, Fig. 2). Isomers with an H-atom at the α position (tertiary C atom) did not serve as growth substrate. Their cometabolic degradation required the presence of an α-quaternary NP isomer. The transformation of the α-tertiary isomers preserved the attachment of the NP group to the aromatic ring and led to the generation of corresponding NP hydroquinones or benzoquinones (pathway B, Fig. 2). The cometabolic formation of nonylbenzoquinones is of particular environmental concern because of the high toxic potential of quinones (Monks et al. 1992). Li et al. (2011) tested the degradation of chlorinated and brominated NP by the strain S. xenophagum Bayram at initial concentrations of 1 and 20 mg L−1 of the halogenated compounds. Deviating from the transformation of the chlorinated compounds, the degradation of the monobromo and dibromo isomers strongly depended on the availability of nonhalogenated NP acting as growth substrate. A repeated addition of NP to microbial cultures exposed to the brominated isomers stimulated the degradation of the latter which would cease otherwise. These findings underline the significance of a former study by Fujita and Reinhard (1997) who stated a cometabolic transformation of brominated octylphenoxyacetic acid to brominated octylphenol and a compound tentatively identified as 2-aminomethoxy-3-bromo-5-(1,1,3,3-tetramethylbutyl)phenol. Some studies were directed toward the NP degradation by NAS. Torres-Bojorges and Buitrón (2012) used nitrifying biomass enriched from activated sludge of a municipal WWTP. According to respirometric tests, the sludge contained less than 0.1 % of heterotrophic bacteria at the end of a 150day acclimatization period. It was found that NAS, fed with a mineral salt medium, was superior in degrading technical NP, supplied in a concentration of 100 μg L−1 at pH 7.5, compared to the genuine activated sludge (75 % of heterotrophic bacteria) and to a 4-chlorophenol-adapted consortia. After 250 h of batch testing, 43 % of NP was degraded by NAS and 18 % by the unmodified activated sludge. A similar approach was followed by Kim et al. (2007) who tested the NP degradation by NAS at higher NP initial concentration (10 mg L−1). The NAS was enriched by means of a sequential batch reactor

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Fig. 2 Metabolism of α-quaternary (a) and α-tertiary (b) NP isomers by strain Sphingonomas xenophagum Bayram. Pathway B proceeds cometabolically in the obligate presence of an α-quaternary NP isomer.

Bold and thin reaction arrows symbolize putatively high and medium rates of reaction, respectively. Reprint with permission from Gabriel et al. (2012)

during more than 4 months. Supplementing the NAS culture with ammonium, the NP elimination reached approximately 90 % after 6 days. When nitrite was added instead of ammonium, the elimination dropped to roughly 30 % indicating that AOB and not nitrite-oxidizing bacteria were responsible for this result.

HRTs of 24 h (aerobic reactor) and of 12 h (anaerobic conditions) were found to be optimal. The optimal mass ratio of the fungicide to the growth substrate glucose was 1:100. The ability of ASMO to degrade the bactericide triclosan was firstly established by Singer et al. (2002). Subsequently, the involvement of AOB was ascertained, providing hints for a cometabolic process (Roh et al. 2009). This hypothesis was further confirmed by Kim et al. (2011) who compared the transformation efficiency of three aerobic sludge species for triclosan. They identified Sphingomonas sp. PH-07 as a microbial strain able to partially degrade triclosan, producing the metabolites hydroxy-triclosan, 4-chlorophenol, and 2,4-dichlorophenol. Chen et al. (2011) ascertained the formation of methyl-triclosan in activated sludge biodegradation tests without additional carbon sources. Recently, Lee et al. (2012) reported the wastewater bacterial isolate Sphingopyxis KCY 1 to degrade the biocide in the obligate presence of glucose, sodium succinate, or sodium acetate. Various chlorinated intermediates of the transformation process were discovered, but finally, the reaction ended up with the formation of unchlorinated products. The degradation was impeded by the addition of 3-fluorocatechol, an inhibitor of meta-cleavage enzymes, e.g., catechol 2,3-dioxygenase, whose presence was determined in cell extracts of the microorganisms. A summary of the various triclosan transformation products is provided in Fig. 3.

Pesticides and biocides Data on the cometabolic degradation of pesticides and biocides by activated sludge microorganisms (ASMO) are very scarce, but the pioneering work on cometabolism in sewage dealt with the four herbicides trifluralin, profluralin, fluchloralin, and nitrofen (Jacobson et al. 1980). Gosh and Philip (2004) studied the transformation of atrazine by an anaerobic-mixed microbial culture in the concentration range of 0.5–15 mg L−1 with and without dextrose as external carbon source. The tests were performed in a lab-scale reactor operated in the sequential batch mode, maintaining a hydraulic retention time (HRT) of about 5 days. The atrazine degradation (first-order reaction rate) was about 20 times faster with the addition of dextrose than without it. For the structurally related triazole fungicide triadimenol, both aerobic and anaerobic treatments led to the degradation of the compound at initial concentrations between 1 and 25 mg L−1 (Shawaqfeh 2010). In laboratory experiments,

Appl Microbiol Biotechnol Fig. 3 Transformation products of triclosan according to Kim et al. (2011) and Chen et al. (2011) (dechlorinated dihydroxytriclosan not shown)

Nitro and nitroso compounds Due to various industrial and commercial sources, nitrosamines (NSA) are frequently detected in wastewater, albeit at low concentrations (sum of NSA typically ≤0.2 μg L−1; Krauss et al. 2009). Various studies revealed NSA degradation in environmental media under aerobic and anaerobic conditions. Sedlak et al. (2005) confirmed NSA degradation by ASMO with greatly varying elimination efficiencies. Since no microorganisms have been isolated so far

that can utilize NSA as sole substrates for growth, a cometabolic degradation process is assumed (Sharp et al. 2 0 0 5; F o u r ni e r e t a l . 2 0 06 ) . Va r i o u s m i c r o b i a l monooxygenases are able to attack NSA (Sharp et al. 2005, 2007). The specific role of AMO remains however unclear. Except for N-nitrosomorpholine, no significant differences were reported between the removal efficiencies under nitrifying and non-nitrifying conditions for five NSA (Krauss et al. 2009) The same authors reported a significant depression of the removal efficiency for nitrosamines when their

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concentrations were below approximately 15 ng L−1. It was assumed that substrate competition in the cometabolic degradation explains this effect. Zhao and Ward (2000) demonstrated the cometabolic transformation of nitrobenzene by a strain of Pseudomonas putida (2NP8) which was isolated from AS. This degradation activity was induced by bacterial growth on 3-nitrophenol but not on 2-nitrophenol. The formation of the metabolites ammonia, nitrosobenzene, and hydroxylaminobenzene indicated the involvement of the enzyme 3-nitrophenol nitroreductase in the degradation process. According to Kroger et al. (2004), 2,4,6-trinitrotoluene (TNT) was cometabolically converted into aminodinitrotoluenes by ASMO under aerobic conditions, while mixtures of aminodiand mononitrotoluenes were generated under anaerobic conditions.

Miscellaneous compounds Alkyl and alkylaryl ethers Hernandez-Perez et al. (2001) reported the cometabolic degradation of methyl-t-butyl ether and of t-amyl methyl ether by Gordonia terrae strains IFP2001 and IFP2007, isolated from AS. Formate was identified as degradation product. The authors supposed the involvement of a monooxygenase in the transformation process. Sei et al. (2010) demonstrated the metabolization of 1,4-dioxane by an AS sample in the presence of tetrahydrofuran. The metabolization efficiency increased with repeated exposition to 1,4-dioxane. The cometabolic degradation of 1,4-dioxane in the presence of tetrahydrofuran as growth substrate by Pseudonocardia sp. strain ENV478 was also reported by Vainberg et al. (2006) and Masuda et al. (2012). The responsible enzyme was a THF-inducible monooxygenase. The biotransformation of several alkyl and alkylaryl ethers by various strains of Rhodococcus sp. and Gordonia sp. was investigated by Kim et al. (2008). The strains Rhodococcus sp. DEOB100 and Gordonia sp. DEOB200 were capable of transforming 1,3-dialkoxybenzenes and 1,4dialkoxybenzenes to the corresponding monoalkoxy phenols in the presence of ethanol. The degradation rates decreased in the order 1,4-diethoxybenzene>1,4-dimethoxybenzene>1,3dimethoxybenzene, while 1,2-dimethoxybenzene was not degraded. Phthalates Yu et al. (2009) ascertained the transformation of dimethyl phthalate with glucose as primary substrate in an SBR, applying the aerobic granulation technique. Depending on organic load and sludge settling time, removal efficiencies >90 % were achieved after 75 days of operation. The degradation kinetics was adequately described using the Haldane model. Di-2-ethylhexyl phthalate and other phthalate diesters were cometabolically degraded by Rhodococcus rhodochrous releasing the corresponding monoester. Biodegradation rates

were found to be strongly influenced by steric hindrances and compound solubility. The responsible esterase, associated with the bacterial cell wall, was not induced by specific substrates and was detected at all stages of microbial growth (Sauvageau et al. 2009). Bisphenol A and tetrabromobisphenol A (TBBPA) The importance of nitrifying bacteria for the degradation of bisphenol A was elucidated by Kim et al. (2007). They enriched nitrifying sludge from AS with low nitrifying activity by means of a SBR for more than 4 months. Parallel tests with NAS supplemented with either ammonium or nitrite as nitrogen source revealed a superior bisphenol A elimination by AOB. Applying various inhibitors of microbial activity (ATU and Hg2SO4) more than 60 % of the total bisphenol A elimination could be attributed to the autotrophic bacteria, whereas approximately equal proportions of 20 % belonged to degradation by heterotrophs and to sludge adsorption. TBBPA is the brominated flame retardant with the highest production volume at present. Anaerobic sludge has a certain potential to degrade TBBPA, depending on its microbial composition (Peng et al. 2012). Applying an ASC enriched with the TBBPA degrader Comamonas sp. strain JXS-2-02, the TBBPA elimination rate was significantly increased (Peng et al. 2013). Batch tests with an enriched anaerobic culture evidenced a strong dependence of the first-order degradation rate on the type of growth substrate (Peng and Jia 2013). At equal initial concentrations of TBBPA and the growth substrate (each 0.5 mg L−1) degradation rates >80 % within 10 days were attained with sodium formate, sodium acetate, and glucose. With TBBPA as sole C source, its transformation degree was below 30 % and abiotic processes might have been involved, especially adsorption to the activated sludge (elimination in sterile control approximately 15 %). The simultaneous supply of the microbial culture with easily bioavailable C and N sources, i.e., formate and yeast extract, further improved the TBBPA degradability. Nevertheless, due to the biomass growth, the efficiency of the adsorptive elimination might had increased also. 4,4’-Diaminostilbene-2,2’-disulfonic acid (DSDA) DSDA, an intermediate of the production of whitening agents, dyes, and fungicides is present in related industrial wastewaters and, at trace concentrations, in communal wastewaters. By means of a lab-scale aerobic SBR reactor, filled with DSDA-acclimated activated sludge from a secondary sedimentation tank of a WWTP, DSDA removal efficiencies for highly loaded batches (DSDA concentrations between 100 and 500 mg L−1) were investigated at various operation conditions (Yan et al. 2010). The addition of glucose as growth substrate significantly increased the process efficiency. The COD removal correlated with the fraction of the dehydrogenase activity stemming from the substrate conversion.

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Iopromide (IOPr) IOPr is an iodinated X-ray contrast medium which has been identified to enter urban water cycles (Ratola et al. 2012). Principally, this compound is degradable by NAS (Batt et al. 2006) and anaerobes (Lecouturier et al. 2008). Recently, the IOPr-degrading bacterium strain Pseudomonas sp. I-24 was isolated from AS of a WWTP in Shanghai (Liu et al. 2013a). The IOPr degradation efficiency (test concentration 30 mg L−1) of the isolated culture strongly depended on the type of added natural substrate with starch as most and peptone as least efficient substrates. In the presence of starch, more than 90 % of IOPr were transformed within 4 days, whereas roughly 90 % persisted, if peptone was added. A similar ratio was found for the activities of crude extracts of the IOPr-degrading enzyme(s) not being further characterized, expressed at different substrate feeding. AS samples enriched with the isolated strain exhibited enhanced IOPr degradation as well but on a lower level than the pure culture. The degree of the removal of the total amount of oxidizable organic compounds (COD) in the wastewater was not influenced by the sludge abundance of the IOPr-degrading bacterium.

Knowledge gaps, methodical trends, and research perspectives Unambiguity and scope of cometabolic process identification The majority of studies were performed with pure or enriched microbial cultures. Often, optimized (in terms of substrate turnover) process conditions were adjusted. It remains unproven whether the determined metabolic activity and pathways are transferable to mixed consortia under realistic conditions. Avoiding high analytical efforts, pollutant degradation experiments are often carried out at concentrations that surpass those of monitored raw sewage by one or several orders of magnitude. More often, the pollutants are added to growth media but not spiked into real sewage and offered as such. If real wastewater is used, the main DOC fractions, i.e., carbohydrates, peptides, lipids, and organic acids, are not determined despite of the fact that the type of growth substrate might have a decisive influence on turnover rates and on the nature of formed transformation products. Even in tests without sewage, different growth substrates are seldom supplied. Except the common differentiation between the groups of heterotrophic and autotrophic bacteria, a comprehensive analysis of the microbial community structure, which would contribute to a rationalization of differences in their cometabolic activities, is missing in contrast to current developments in the field of biotechnological generation of renewable energy and sludge digestion (Regueiro et al. 2012; Van Wonterghem et al. 2014). Another critical point is the negligence of transformation products. One of the first studies in this field, conducted by

Alexander and coworkers (Jacobson et al. 1980), has addressed that topic, but many of the following did not. With the increasing awareness of the importance of degradation products for the assessment of the environmental impact and (eco-)toxic effects of synthetic chemicals, the total number of investigations pointing to that issue is increasing and so the (rather small) portion that relates transformation products to cometabolic processes during sewage treatment. Some of these studies greatly broadened our understanding of the chemical mechanisms of transformation reactions (Gabriel et al. 2005b), and they discovered correlations between the structure of growth substrates, the nature of induced enzymes, and the type of generated transformation products (Khunjar et al. 2011). Molecular process level Recalling the statement of Criddle (1993) that “cometabolism results from the lack of specifity of enzymes and cofactors”, it is surprising to notice the huge gap between importance and research efforts directed toward that point. As Table 1 reveals, this topic has attracted increasing attention within the last 5 years. Mostly oxidoreductases were surveyed, but the potential contributions of other enzyme classes, i.e., hydrolases, are largely unexplored. Except for studies with pure microbial cultures, the origin of the enzymes is often not clear. It has recently been shown that the unproven affiliation of enzymes to groups of microbial species, i.e., monooxygenases to AOB, might not be justified (Helbling et al. 2012). The reliability of enzyme identification is often low, as mostly indirect enzyme measurements were carried out. Selective enzyme assays—as far as those exist—were seldom applied. The isolation or even fractionation of targeted enzymes from ASC to confirm their hypothesized participation in cometabolic transformation processes is de facto not addressed. In cases where extracellular enzymes are scrutinized, which are mainly associated with sludge flocs, research is needed to clarify the dependence of their accessibility, selectivity, activity, and lifetime from their physical and chemical associations or entrapment with or in flocs. Mazzei et al. (2013) have recently evidenced that the activity of alkaline phosphatase in soils can be significantly inhibited due to hostguest interactions with humic superstructures. Sewage constituents, e.g., metal ions, might also modify enzyme activities. Another question deals with the induction of biosynthesis of the relevant enzymes. Usually, structurally analogous compounds are recognized as inductors. Nevertheless, cometabolic reactions also cope with structurally nonanalogous compounds (Brandt et al. 2003). The occurrence and specific mode of enzyme induction might also depend on the evolutionary state of the degradation pathway linking

Appl Microbiol Biotechnol Table 1 Microbial enzymes potentially participating in cometabolic pollutant degradation Micropollutant

Enzyme(s)

Enzyme detection technique Biomass

Reference

Benzafibrate

Amidases

None (assumed)a

ASC3

Atenolol Bisphenol A Dialkylether (t-amyl methyl, methyl-t-butyl) Dioxane

AMO4 AMO Monooxygenase

Inhibition (ATU)2 Inhibition (ATU) Indirect (detection of cytochrome p-450) Gene expression analysis

NAS6 Nitrosomonas europaea Gordonia terrae strain ifp 2001 Pseudonocardia sp. Strain env478 NAS

Quintana et al. (2005) Helbling et al. (2010b) Sathyamoorthy et al. (2013) Roh et al. (2009) Hernandez-Perez et al. (2001)

17α-ethyl-estradiol (EE2)

THF5-inducible monooxygenase AMO

Ee2

AMO

Ee2

Toluene dioxygenase, catechol 1,2-dioxygenase, catechol-2,3-dioxygenase Dioxygenases 3-Nitrophenol nitroreductase

Ketoprofen Nitrobenzene N-nitrosodimethylamine (NDMA) Sulfamethoxazole (SMX)

Triclosan

Various monooxygenases Arylamine-N-acetyltransferase Amidases Urethanase N-acetyl-phenylethylamine-hydrolase AMO 2,3-Dioxygenases

Enzyme extraction, chromatographic separation, calc. of NADH consumption Inhibition (ATU)

NAS

Specific assays

ASC

None (assumed) None (derived from transformation products) Gene expression analysis

ASC Pseudomonas putida strain 2np8 Rhodococcus sp. strain RHA1 Rhodococcus equi

None (assumed) None (assumed) None (assumed) None (assumed)

Vainberg et al. (2006) Masuda et al. (2012) Yi and Harper (2007)

Shi et al. (2004) Maeng et al. (2013) Khunjar et al. (2011) Khunjar et al. (2011)

Quintana et al. (2005) Zhao and Ward (2000) Sharp et al. (2005, 2007) Larcher and Yargeau (2011)

Inhibition (ATU) Nitrosomonas europaea Roh et al. (2009) Inhibition (3-fluorocatechol) Sphingopyxis strain KCY 1 Lee et al. (2012)

ATU allylthiourea, ASC activated sludge consortium, AMO ammonia monooxygenase, THF tetrahydrofuran, NAS nitrifying activated sludge a

Deduced from metabolic logic and experimental indications

distinct enzymes into a metabolic process chain. Recent studies observed the generation of new pollutant degradation pathways in their “status nascendi” (Copley et al. 2012; Aylward et al. 2013). In this developmental stage, where new enzymatic competences are acquired by (mega)plasmid transfer to a considerable extent, the role of specific chemical inductors seems not to be decisive. In this context, the concept of promiscuous enzymes or enzyme promiscuity might be very promising to deepen the understanding of cometabolic reactions and to relocate research perspectives. A central aspect is to decipher whether and how micropollutants influence the evolution of the promiscuous capabilities of enzymes. Finally, current research on pollutant transformation intends to unravel the “metabolic logic” behind those processes and to use metabolic rules for the target analysis of unknown transformation products (Kern et al. 2010; Helbling et al. 2010a, b). By means of computer-based biotransformation pathway prediction systems, e.g., UM-BBD (Gao et al. 2011), PathPred (Moriya et al. 2010), and CATABOL

(Dimitrov et al. 2007), probable structures of transformation products are generated. Here, the crucial point is that these databases do not distinguish between metabolic and cometabolic transformation processes. Thus, cometabolic pathways, involved enzymes, and formed transformation products are either not covered or cannot be specifically retrieved. One alternative, to classify promiscuous enzyme activities according to the Enzyme Commission (EC), numbering system is not generally accepted yet (Khersonsky and Tawfik 2010). Cometabolism and the community level Originally, the term “cometabolism” was coined to categorize biotransformation processes on the single species level. The simplest way to transfer its meaning on the interspecies level is provided by the term “commensalism”, describing a type of interaction, where one species benefits from the other, e.g., consuming cometabolically generated substrates, whereas the second one, the cometabolically active species, has neither advantages nor

Appl Microbiol Biotechnol

disadvantages. Nowadays, we are aware that microbial communities create “metabolic networks” or a “supermetabolome” with an unique metabolic quality based on its capability to orchestrate a multitude of metabolic pathways and to evolve new metabolic tools, including new (or more promiscuous) enzymes. As a consequence, it is justified to state “that the biomass from each WWTP is unique and has a different set of micropollutant biotransformation capabilities” (Helbling et al. 2012). From this perspective, some but surely not all cometabolic reactions might lose their fortuitousness. It seems to be a great challenge to delineate the metabolic network of activated sludge communities and to discover the functioning of cometabolic processes in the dynamic assembly of established and newly evolving biotransformation pathways.

References Alvarez-Cohen L, Speitel GE (2001) Kinetics of aerobic cometabolism of chlorinated solvents. Biodegradation 12:105–126 Aylward FO, McDonald BR, Adams SM, Valenzuela A, Schmidt RA, Goodwin LA, Woyke T, Currie CR, Suen G, Poulsen M (2013) Comparison of 26 Sphingomonad genomes reveals diverse environmental adaptations and biodegradative capabilities. Appl Environ Microbiol 79:3724–3733 Ambrosio ST, Campos-Takaki GM (2004) Decolorization of reactive azo dyes by Cunninghamella elegans UCP 542 under co-metabolic conditions. Bioresour Technol 91:69–75 Andersen HR, Siegrist H, Halling-Sorensen B, Ternes TA (2003) Fate of estrogens in municipal sewage treatment plant. Environ Sci Technol 37:4021–4026 Babtie A, Tokuriki N, Hollfelder F (2010) What makes an enzyme promiscuous? Curr Opin Chem Biol 14:200–207 Batt AL, Kim S, Aga DS (2006) Enhanced biodegradation of iopromide and trimethoprim in nitrifying activated sludge. Environ Sci Technol 40:7367–7373 Bouju H, Ricken B, Beffa T, Corvini PF, Kolvenbach BA (2012) Isolation of bacterial strains capable of sulfamethoxazole mineralization from an acclimated membrane bioreactor. Appl Environ Microbiol 78: 277–279 Brandt BW, van Leeuwen IMM, Kooijman SALM (2003) A general model for multiple substrate biodegradation. Application to cometabolism of structurally non-analogous compounds. Water Res 37:4843–4854 Cajthawl T, Kresinova Z, Svobodova K, Sigler K, Rezanka T (2009) Microbial transformation of synthetic estrogen 17α-ethinylestradiol. Environ Pollut 157:3325–3335 Cao Q, Yu Q, Donnell DW (2008) Degradation rate constants of steroids in sewage works and receiving water. Environ Technol 29:1321–1330 Çeçen F, Kocameni BA, Aktas Ö (2009) Metabolic and co-metabolic degradation of industrially important chlorinated organics under aerobic conditions. In: Fatta-Kassimos D et al (eds) Xenobiotics in the urban water cycle, Series “Environmental Pollution”, vol 16. Springer, Heidelberg, pp 161–178 Chang H-L, Alvarez-Cohen L (1995) Model for the cometabolic biodegradation of chlorinated organics. Environ Sci Technol 29:2357–2367 Chen X, Nielsen JL, Furgal K, Liu Y, Lolas IB, Bester K (2011) Biodegradation of triclosan and formation of methyl-triclosan in activated sludge under aerobic conditions. Chemosphere 84:452– 456

Clara M, Strenn B, Saracevic E, Kreuzinger N (2004) Adsorption of bisphenol-A, 17β-estradiol and 17α-ethinylestradiol to sewage sludge. Chemosphere 56:843–851 Clouzot L, Marrot B, Doumenq P, Roche N (2008) 17α-ethinylestradiol: an endocrine disruptor of great concern. Analytical methods and removal processes applied to water purification. A review. Environ Prog 27:383–396 Collado N, Buttiglieri G, Marti E, Ferrando-Climent L, Rodriguez-Mozaz S, Barcelo D, Comas J, Rodriguez-Roda I (2013) Effects on activated sludge bacterial community exposed to sulfamethoxazole. Chemosphere 93:99–106 Copley SD, Rokicki J, Turner P, Daligault H, Nolan M, Land M (2012) The whole genom sequence of Sphingobium chlorophenolicum L-1: insights into the evolution of the pentachlorophenol degradation pathway. Genome Biol Evol 4:184–198 Corvini PFX, Schäffer A, Schlosser D (2006) Microbial degradation of nonylphenol and other alkylphenols—our evolving view. Appl Microbiol Biotechnol 72:223–243 Criddle CS (1993) The kinetics of cometabolism. Biotechnol Bioeng 72: 1048–1056 Dalton H, Stirling DI (1982) Co-metabolism. Phil Trans R Soc London B 297:481–496 Daniel JA, Chris MY, Michael RH (2001) Molecular and cellular fundamentals of aerobic cometabolism of trichloroethylene. Biodegradation 12:81–103 De Gusseme B, Pycke B, Hennebel T, Marcoen A, Vlaeminck SE, Noppe H, Boon N, Verstraete W (2009) Biological removal of 17αethinylestradiol by a nitrifier enrichment culture in a membrane bioreactor. Water Res 43:2493–2503 Delgadillo-Mirquez L, Lardon L, Steyer J-P, Patureau D (2011) A new dynamic model for bioavailability and cometabolism of micropollutants during anaerobic digestion. Water Res 45:4511–4521 Dimitrov S, Pavlov T, Nedelcheva D, Reuschenbach P, Silvani M, Bias R, Comber M, Low L, Lee C, Parkerton T, Mekenyan O (2007) A kinetic model for predicting biodegradation. SAR QSAR Environ Res 18:443–457 Dytczak MA, Londry KL, Oleszkiewicz JA (2008) Biotransformation of estrogens in nitrifying activated sludge under aerobic and alternating anoxic/aerobic conditions. Water Environ Res 80:47–52 Eganhouse R, Pontolillo J, Gaines R, Frysinger G, Gabriel FLP, Kohler H-PE, Giger W, Barber L (2009) Isomer-specific determination of 4nonylphenols using comprehensive two-dimensional gas chromatography/time-of-flight mass spectrometry. Environ Sci Technol 43:9306–9313 Elias M, Dupuy J, Merone L, Mandrich L, Porzio E (2008) Structural basis for natural lactonase and promiscuous phosphotriesterase activities. J Mol Biol 379:1017–1028 Estrada-Arriaga EB, Mijaylova NP (2010) A comparison of biodegradation kinetic models applied to estrogen removal with nitrifying activated sludge. Water Sci Technol 62:2183–2189 Evangelista S, Cooper DG, Yargeau V (2010) The effect of structure and a secondary carbon source on the microbial degradation of chlorophenoxy acids. Chemosphere 79:1084–1088 Fischer K, Fries E, Körner W, Schmalz C, Zwiener C (2012) New developments in the trace analysis of organic water pollutants. Appl Microbiol Biotechnol 94:11–28 Forrez I, Carballa M, Noppe H, De Brabander H, Boon N, Verstraete W (2009) Influence of manganese and ammonium oxidation on the removal of 17α-ethinylestradiol (EE2). Water Res 43:77–86 Fournier D, Hawari J, Streger SH, McClay K, Hatzinger PB (2006) Biotransformation of N-nitrosodimethylamine by Pseudomonas mendocina KR1. Appl Environ Microbiol 72:6693–6698 Fujita Y, Reinhard M (1997) Identification of metabolites from the biological transformation of the nonionic surfactant residue octylphenoxyacetic acid and its brominated analog. Environ Sci Technol 31:1518–1524

Appl Microbiol Biotechnol Gabriel FLP, Giger W, Guenther K, Kohler H-PE (2005a) Differential degradation of nonylphenol isomers by Sphingonomas xenophaga Bayram. Appl Environ Microbiol 71:1123–1129 Gabriel FLP, Heidlberger A, Rentsch D, Giger W, Guenther K, Kohler HPE (2005b) A novel metabolic pathway for degradation of 4nonylphenol environmental contaminants by Sphingonomas xenophaga Bayram. J Biol Chem 280:15525–15533 Gabriel FLP, Mora MA, Kolvenbach BA, Corvini PFX, Kohler H-PE (2012) Formation of toxic 2-nonyl-p-benzoquinones from α-tertiary 4-nonylphenol isomers during microbial metabolism of technical nonylphenol. Environ Sci Technol 46:5979–5987 Gao J, Ellis LB, Wackett LP (2011) The University of Minnesota pathway prediction system: multi-level prediction and visualization. Nucleic Acids Res 39:W406–W411 Gaulke LS, Strand SE, Kalhorn TF, Stensel HD (2008) 17αethinylestradiol transformation via abiotic nitration in the presence of ammonia oxidizing bacteria. Environ Sci Technol 42:7622–7627 Gauthier H, Yargeau V, Cooper DG (2010) Biodegradation of pharmaceuticals by Rhodococcus rhodochrous and Aspergillus niger by cometabolism. Sci Total Environ 408:1701–1706 Giger W, Gabriel FLP, Jonkers N, Wettstein FE, Kohler H-PE (2009) Environmental fate of phenolic endocrine disruptors: field and laboratory studies. Phil Trans R Soc A 367:3941–3963 Gosh PK, Philip L (2004) Atrazine degradation in anaerobic environment by a mixed microbial consortium. Water Res 38:2277–2284 Gutendorf B, Westendorf J (2001) Comparison of an array of in vitro assays for the assessment of the estrogenic potential of natural and synthetic estrogens, phytoestrogens and xenoestrogens. Toxicology 166:79–89 Harper WF Jr, Floyd-Smith T, Yi T (2008) Chemical processes during biological wastewater treatment. In: Aga DS (ed) Fate of pharmaceuticals in the environment and in water treatment systems. CRC Press, Boca Raton, pp 363–382 Helbling DE, Hollender J, Kohler H-P, Singer H, Fenner K (2010a) Highthroughput identification of microbial transformation products of organic micropollutants. Environ Sci Technol 44:6621–6627 Helbling DE, Hollender J, Kohler H-P, Singer H, Fenner K (2010b) Structure-based interpretation of biotransformation pathways of amide-containing compounds in sludge-seeded bioreactors. Environ Sci Technol 44:6628–6635 Helbling DE, Johnson DR, Honti M, Fenner K (2012) Micropollutant biotransformation kinetics associate with WWTP process parameters and microbial community characteristics. Environ Sci Technol 46:10579–10588 Hernandez-Perez G, Fayolle F, Vandecasteele J-P (2001) Biodegradation of ethyl t-butyl ether (ETBE), methyl t-butyl ether (MTBE), and tamyl methylether (TAME) by Gordonia terrae. Appl Microbiol Biotechnol 55:117–121 Hult K, Berglund P (2007) Enzyme promiscuity: mechanisms and applications. Trends Biotechnol 25:231–238 Ieda T, Horii Y, Petrick G, Yamashita N, Ochiai N, Kannan K (2005) Analysis of nonylphenol isomers in a technical mixture and in water by comprehensive two-dimensional gas chromatography mass spectrometry. Environ Sci Technol 39:7202–7207 Jacobson SN, O’Mara NL, Alexander M (1980) Evidence for cometabolism in sewage. Appl Environ Microbiol 40:917–921 Joss A, Anderson H, Ternes T, Richle PR, Siegrist H (2004) Removal of estrogens in municipal wastewater treatment under aerobic and anaerobic conditions: consequences for plant optimization. Environ Sci Technol 38:3047–3055 Kern S, Baumgartner R, Helbling DE, Hollender J, Singer H, Loos MJ, Schwarzenbach RP, Fenner K (2010) A fixed procedure for assessing the formation of biotransformation products of pharmaceuticals and biocides during activated sludge treatment. J Environ Monit 12:2100–2111 Khersonsky O, Tawfik DS (2010) Enzyme promiscuity: a mechanistic and evolutionary perspective. Annu Rev Biochem 79:471–505

Khunjar WO, Mackintosh SA, Skotnicka-Pitak J, Baik S, Aga DS, Love NG (2011) Elucidating the relative roles of ammonia oxidizing and heterotrophic bacteria during the biotransformation of 17αethinylestradiol and trimethoprim. Environ Sci Technol 45:3605– 3612 Kim JY, Ryu K, Kim EJ, Choe WS, Cha GC, Yoo I-K (2007) Degradation of bisphenol A and nonylphenol by nitrifying activated sludge. Process Biochem 42:1470–1474 Kim Y-H, Cha C-J, Engesser K-H, Kim S-J (2008) Degradation of various alkyl ethers by alkyl ether-degrading Actinobacteria isolated from activated sludge of a mixed wastewater treatment. Chemosphere 73:1442–1447 Kim YM, Murugesan K, Schmidt S, Bokare V, Jeon JR, Kim EJ, Cnang YS (2011) Triclosan susceptibility and co-metabolism—a comparison for three aerobic pollutant-degrading bacteria. Bioresour Technol 102:2206–2212 Kohler H-PE, Gabriel FLP, Giger W (2008) Ipso-substitution: a novel pathway for microbial metabolism of endocrine-disrupting 4nonylphenols, 4-alkoxyphenols, and bisphenol A. Chimica 62: 358–363 Kraigher B, Kosjek T, Heath E, Kompare B, Mandic-Mulec I (2008) Influence of pharmaceutical residues on the structure of activated sludge bacterial communities in wastewater treatment bioreactors. Water Res 42:4578–4588 Krauss M, Longrée P, Dorusch F, Ort C, Hollender J (2009) Occurrence and removal of N-nitrosamines in wastewater treatment plants. Water Res 43:4381–4391 Kroger M, Schumacher ME, Risse H, Fels G (2004) Biological reduction of TNT as part of combined biological-chemical procedure for mineralization. Biodegradation 15:241–248 Larcher S, Yargeau V (2011) Biodegradation of sulfamethoxazole by individual and mixed bacteria. Appl Microbiol Biotechnol 91:211– 218 Lecouturier D, Rochex A, Lebeault JM (2008) The mineralization of 5amino-2,4,6-triiodoisophthalic acid by a two-stage fixed-bed reactor. Water Res 42:2491–2498 Lee DG, Zhao F, Rezenom YH, Russell DH, Chu KH (2012) Biodegradation of triclosan by a wastewater microorganism. Water Res 46:4426–4434 Li Y, Montgomery-Brown J, Reinhard M (2011) Biotransformation of halogenated nonylphenols with Sphingobium xenophagum Bayram and a nonylphenol-degrading soil-enrichment culture. Arch Environ Contam Toxicol 60:212–219 Likitmongkonsakun N, Limpiyakorn T (2013) Effect of real wastewater on biotransformation of 17α-ethinylestradiol by ammonia-oxidizing bacteria in nitrifiying activated sludge. Int J Environ Earth Sci Eng 7:82–86 Limpiyakorn T, Homklin S, Ong SK (2011) Fate of estrogens and estrogenic potentials in sewage systems. Crit Rev Environ Sci Technol 41:1231–1270 Liu Y, Hu J, Xu B, He J, Gao P, Liu K, Xue G, Ognier S (2013a) Isolation and identification of a iopromide-degrading strain and its application in an A2/O-system. Bioresour Technol 134:36–42 Liu Z, Dai Y, Huan Y, Sun L, Zhou Q, Zhang W, Sang Q, Wei H, Yuan S (2013b) Different utilizable substrates have different effects on cometabolic fate of imidacloprid in Stenotrophomonas maltophilia. Appl Microbiol Biotechnol 97:6537–6547 Lozano N, Rice CP, Ramirez M, Torrents A (2013) Fate of triclocarbon, triclosan and methyltriclosan during wastewater and biosolids treatment processes. Water Res 47:4519–4527 Maeng SK, Choi BG, Lee KT, Song KG (2013) Influences of solid retention time, nitrification and microbial activity on the attenuation of pharmaceuticals and estrogens in membrane bioreactors. Water Res 47:3151–3162 Majewsky M, Gallé T, Yargeau V, Fischer K (2011) Active heterotrophic biomass and sludge retention time (SRT) as determining factors for

Appl Microbiol Biotechnol biodegradation kinetics of pharmaceuticals in activated sludge. Bioresour Technol 102:7415–7421 Masuda H, McClay K, Steffan RJ, Zylstra GJ (2012) Biodegradation of tetrahydrofuran and 1,4-dioxane by soluble diiron monooxygenase in Pseudonocardia sp. strain ENV478. J Mol Microbiol Biotechnol 22:312–316 Mazzei P, Oschkinat H, Piccolo A (2013) Reduced activity of alkaline phosphatase due to host-guest interactions with humic superstructures. Chemosphere 93:1972–1979 Mes TZDD, Kujawa-Roeleveld K, Zeeman G, Lettinga G (2008) Anaerobic biodegradation of estrogens —hard to digest. Water Sci Technol 57:1177–1182 Monks TJ, Hanzlik RP, Cohen GM, Ross D, Graham DG (1992) Quinone chemistry and toxicity. Toxicol Appl Pharmacol 112:2–6 Moriya Y, Shigemizu D, Hattori M, Tokimatsu T, Kotera M, Goto S, Kanehlsa M (2010) PathPred: an enzyme-catalyzed metabolic pathway prediction server. Nucleic Acids Res 38:W138–W143 Müller E, Schüssler W, Horn H, Lemmer H (2013) Aerobic biodegradation of the sulfonamide antibiotic sulfamethoxazole by activated sludge applied as co-substrate and sole carbon and nitrogen source. Chemosphere 92:969–978 Nalli S, Cooper DG, Nicell JA (2006) Interaction of metabolites with R. rhodochrous during the biodegradation of di-ester plasticizers. Chemosphere 65:1510–1517 Ottmar KJ, Colosi LM, Smith JA (2012) Fate and transport of atorvastatin and simvastatin drugs during conventional wastewater treatment. Chemosphere 88:1184–1189 Pajot HF, Farina JI, De Figueroa LIC (2011) Evidence on manganese peroxidase and tyrosinase expression during decolorization of textile industry dyes by Trichosporon akiyoshidainum. Int Biodeterior Biodegrad 65:1199–1207 Pauwels B, Noppe H, De Brabander H, Verstraete W (2008) Comparison of steroid hormone concentration in domestic and hospital wastewater treatment plants. J Environ Eng ASCE 134:933–936 Peng XX, Jia XS (2013) Optimization of parameters for anaerobic cometabolic degradation of TBBPA. Bioresour Technol 148:386–393 Peng XX, Zhang ZL, Luo WS, Jia XS (2013) Biodegradation of tetrabromobisphenol A by a novel Comamonas sp. strain, JXS-202, isolated from anaerobic sludge. Bioresour Technol 128:173–179 Peng XX, Zhang ZL, Zhao ZL, Jia XS (2012) 16S ribosomal DNA clone libraries to reveal bacterial diversity in anaerobic reactor-degraded tetrabromobisphenol A. Bioresour Technol 112:75–82 Pholchan P, Jones M, Donnelly T, Sallis PJ (2008) Fate of estrogens during the biological treatment of synthetic wastewater in a nitriteaccumulating sequencing batch reactor. Environ Sci Technol 42: 6141–6147 Plosz BG, Leknes H, Thomas KV (2010) Impacts of competitive inhibition, parent compound formation and partitioning behavior on the removal of antibiotics in municipal wastewater treatment. Environ Sci Technol 44:734–742 Pomiès M, Choubert J-M, Wisniewski C, Coquery M (2013) Modelling of micropollutant removal in biological wastewater treatments: a review. Sci Total Environ 443:733–748 Quintana JB, Weiss S, Reemtsma T (2005) Pathways and metabolites of microbial degradation of selected acidic pharmaceutical and their occurrence in municipal wastewater treated by a membrane bioreactor. Water Res 39:2654–2664 Racz LA, Muller JG, Goel RK (2012) Fate of selected estrogens in two laboratory scale sequencing batch reactors fed with different organic carbon sources under varying solids retention times. Bioresour Technol 110:35–42 Ratola N, Cincinelli A, Alves A, Athanasios K (2012) Occurrence of organic micro contaminants in the wastewater treatment process. A mini review. J Hazard Mater 239–240:1–18 Reemtsma T, Jekel M (eds) (2006) Organic pollutants in the water cycle. Wiley-VCH, Weinheim

Regueiro L, Veiga P, Figueroa M, Alonso-Gutierez J, Stams AJM, Lema JM, Carballa M (2012) Relationship between microbial activity and microbial community structure in six full-scale anaerobic digesters. Microbiol Res 167:581–589 Ren Y-X, Nakano K, Nomura M, Chiba N, Nishimura O (2007) Effects of bacterial activity on estrogen removal in nitrifying activated sludge. Water Res 41:3089–3096 Richardson SD, Ternes TA (2011) Water analysis: emerging contaminants and current issues. Anal Chem 83:4616–4648 Roh H, Subramanya N, Zhao FM, Yu CP, Sandt J, Chu KH (2009) Biodegradation potential of wastewater micropollutants by ammonia-oxidizing bacteria. Chemosphere 77:1084–1089 Sathyamoorthy S, Chandran K, Ramsburg CA (2013) Biodegradation and cometabolic modeling of selected beta blockers during ammonia oxidation. Environ Sci Technol 47:12835–12843 Sauvageau D, Cooper DG, Nicell JA (2009) Relative rates and mechanisms of biodegradation of diester plasticizers mediated by Rhodococcus rhodochrous. Can J Chem Eng 87:499–506 Sayavedra-Soto LA, Gvakharia B, Bottomley PJ, Arp DJ, Dolan ME (2010) Nitrification and degradation of halogenated hydrocarbons— a tenuous balance for ammonia-oxidizing bacteria. Appl Microbiol Biotechnol 86:435–444 Sedlak DL, Deeb RA, Hawley EL, Mitch WA, Durbin TD, Mowbray S, Carr S (2005) Sources and fate of nitrosodimethylamine and its precursors in municipal wastewater treatment plants. Water Environ Res 77:32–39 Sei K, Kakinoki T, Inoue D, Soda S, Fujita M, Ike M (2010) Evaluation of the biodegradation potential of 1,4-dioxane in river, soil and activated sludge samples. Biodegradation 21:585–591 Sharp JO, Sales CM, LeBLanc JC, Liu J, Wood TK, Eltis LD, Mohn WW, Alvarez-Cohen L (2007) An inducible propane monooxygenase is responsible for N-nitrosodimethylamine degradation by Rhodococcus sp. strain RHA1. Appl Environ Microbiol 73:6930–6938 Sharp JO, Wood TK, Alvarez-Cohen L (2005) Aerobic biodegradation of N-nitrosodimethylamine (NDMA) by axenic bacterial strains. Biotechnol Bioeng 89:608–618 Shawaqfeh AT (2010) Removal of pesticides from water using anaerobicaerobic biological treatment. Chin J Chem Eng 18:672–680 Shi J, Fujisawa S, Nakai S, Hosomi M (2004) Biodegradation of natural and synthetic estrogens by nitrifying activated sludge and ammonia oxidizing bacterium Nitrosomonas europaea. Water Res 38:2323–2330 Singer H, Muller S, Tixier C, Pillonel L (2002) Triclosan: occurrence and fate of a widely used biocide in the aquatic environment: field measurements in wastewater treatment plants, surface waters, and lake sediments. Environ Sci Technol 36:4998–5004 Sugumar WR, Sadanandan S (2010) Combined anaerobic–aerobic bacterial degradation of dyes. E-J Chem 7:739–744 Ternes A, Joss A (eds) (2006) Human pharmaceuticals, hormones and fragrances. IWA Publishing, London Torres-Bojorges AX, Buitrón G (2012) Biodegradation of nonylphenols using nitrifying sludge, 4-chlorophenol-adapted consortia and activated sludge in liquid and solid phases. Environ Technol 33:1727–1737 Tran NH, Urase T, Kusakabe O (2009) The characteristics of enriched nitrifier culture in the degradation of selected pharmaceutically active compounds. J Hazard Mater 171:1051–1057 Tran NH, Urase T, Ngo HH, Hu J, Ong SL (2013) Insight into metabolic and cometabolic activities of autotrophic and heterotrophic microorganisms in the biodegradation of emerging trace organic contaminants. Bioresour Technol 146:721–731 Urase T, Kikuta T (2005) Separate estimation of adsorption and degradation of pharmaceutical substances and estrogens in the activated sludge process. Water Res 39:1289–1300 Vainberg S, McClay K, Masuda H, Root D, Condee C, Zylstra GJ, Steffan RJ (2006) Biodegradation of ether pollutants by Pseudonocardia sp. strain ENV478. Appl Environ Microbiol 72: 5218–5224

Appl Microbiol Biotechnol Van Wonterghem I, Jensen PD, Ho DP, Batstone DJ, Tyson GW (2014) Linking microbial community structure, interactions and function in anaerobic digesters using new molecular techniques. Curr Opin Biotechnol 27:55–64 Wendlandt KD, Stottmeister U, Helm J, Soltmann B, Jechorek M, Beck M (2010) The potential of methane-oxidizing bacteria for applications in environmental biotechnology. Eng Life Sci 10:87–102 Xia S, Jia R, Feng F, Xie K, Li H, Jing D, Xu X (2012) Effects of solids retention time on antibiotics removal performance and microbial communities in an A/O-MBR process. Bioresour Technol 106:36– 43 Xu K, Harper WF Jr, Zhao D (2008) 17α-ethinylestradiol sorption to activated sludge biomass: thermodynamic properties and reaction mechanism. Water Res 42:3146–3152

Yan Y, Feng L, Sun L, Zhou Q (2010) Integration of a coagulation step in a cometabolism sequencing batch reactor for the treatment of DSD acid wastewater from a reduction process. Desalination 250:167–171 Yi T, Harper WF (2007) The link between nitrification and biodegradation of 17 alpha-ethinylestradiol. Environ Sci Technol 41:4311–4316 Yu C-P, Deeb RA, Chu K-H (2013) Microbial degradation of steroidal estrogens. Chemosphere 91:1225–1235 Yu H-Y, Yao L, Ye Z-F (2009) Aerobic granulation for dimethyl phthalate biodegradation in a sequencing batch reactor. Huanjing Kexue/ Environ Sci 30:2661–2666 Zhao J-S, Ward OP (2000) Cometabolic biotransformation of nitrobenzene by 3-nitrophenol degrading Pseudomonas putida 2NP8. Can J Microbiol 46:643–652 Zhou X, Oleszkiewicz JA (2010) Biodegradation of estrogens in nitrifying activated sludge. Environ Technol 31:1263–1269