Abstract
Triclosan is considered a ubiquitous pollutant and can be detected in a wide range of environmental samples. Triclosan removal by wastewater treatment plants has been largely attributed to biodegradation processes; however, very little is known about the micro-organisms involved. In this study, DNA-based stable isotope probing (DNA-SIP) combined with microautoradiography-fluorescence in situ hybridization (MAR-FISH) was applied to identify active triclosan degraders in an enrichment culture inoculated with activated sludge. Clone library sequences of 16S rRNA genes derived from the heavy DNA fractions of enrichment culture incubated with 13C-labelled triclosan showed a predominant enrichment of a single bacterial clade most closely related to the betaproteobacterial genus Methylobacillus. To verify that members of the genus Methylobacillus were actively utilizing triclosan, a specific probe targeting the Methylobacillus group was designed and applied to the enrichment culture incubated with 14C-labelled triclosan for MAR-FISH. The MAR-FISH results confirmed a positive uptake of carbon from 14C-labelled triclosan by the Methylobacillus. The high representation of Methylobacillus in the 13C-labelled DNA clone library and its observed utilization of 14C-labelled triclosan by MAR-FISH reveal that these micro-organisms are the primary consumers of triclosan in the enrichment culture. The results from this study show that the combination of SIP and MAR-FISH can shed light on the networks of uncultured micro-organisms involved in degradation of organic micro-pollutants.
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The GenBank/EMBL/DDBJ accession numbers for the sequences of the triclosan-degrading culture clones represented in the phylogenetic tree are JX099503–JX099536.
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A supplementary figure is available with the online version of this paper.
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Edited by: E. L. Madsen
Introduction
Triclosan [5-chloro-2-(2,4-dichloro-phenoxy)-phenol] is a synthetic antibacterial compound that inhibits the NADH-dependent enoyl-[acyl-carrier protein] reductase, an essential enzyme involved in the biosynthesis of fatty acids (Heath et al., 1999; McMurry et al., 1998; Regös et al., 1979). As an effective antimicrobial agent, triclosan has been used in a wide range of personal care products, such as toothpaste and soaps, and in consumer products, including textile and plastics (DeSalva et al., 1989; Jones et al., 2000; Schweizer, 2001). Due to its extensive use and persistence, triclosan and some of its derivatives can be detected in different environmental matrices such as wastewaters, surface waters and sediments, and in biological samples, including those from fish, algae, human plasma, urine and breast milk (Balmer et al., 2004; Hovander et al., 2002; Miller et al., 2008; Sánchez-Brunete et al., 2010; Sandborgh-Englund et al., 2006; Wilson et al., 2003; Ye et al., 2008).
Biodegradation of triclosan has been shown by mixed bacterial cultures from activated sludge (Gangadharan Puthiya Veetil et al., 2012; Hay et al., 2001; Stasinakis et al., 2010) and in wastewater treatment plants (WWTPs) (Bester, 2003; Chen et al., 2011; Singer et al., 2002). Due to insufficient removal during wastewater treatment, triclosan has been found in WWTP effluents in concentrations ranging from 1 to 10 µg l−1 (Adolfsson-Erici et al., 2002; Bester, 2003, 2005; Lindström et al., 2002; Singer et al., 2002). Although mass balance assessments have shown that biological treatment contributes to the major removal of triclosan in WWTPs (Bester, 2003; Heidler & Halden, 2007; Singer et al., 2002), little is known about the actual mechanisms or the micro-organisms involved in the degradation process. So far, two wastewater isolates, Sphingomonas sp. strains Rd1 (Hay et al., 2001) and PH-07 (Kim et al., 2011), have been shown to degrade triclosan via co-metabolism. Meade et al. (2001) showed that two soil bacteria, Pseudomonas putida and Alcaligenes xylosoxidans, have a high resistance to triclosan and can utilize it as their sole carbon source, and the nitrifying Nitrosomonas europea has also been shown to biodegrade triclosan (Roh et al., 2009). Recently, several triclosan-degrading strains belonging to the genus Pseudomonas were isolated from aerobic and anaerobic enrichment cultures of activated sludge (Gangadharan Puthiya Veetil et al., 2012). Biodegradation of triclosan has also been reported in fungi (Hundt et al., 2000). However, knowledge based on culture-independent approaches of the identity and ecophysiology of triclosan-degrading bacteria in complex microbial systems is still limited.
Stable-isotope probing (SIP) allows for in situ detection of bacterial communities capable of metabolizing a specific carbon source and thus links function to identity without the need to culture the bacteria involved (Radajewski et al., 2000). SIP approaches have been used to identify various types of environmental pollutant-degrading bacteria, e.g. those able to degrade nonylphenols (Zemb et al., 2012), toluene (Woods et al., 2011) and phenols (Manefield et al., 2007).
The aim of this study was to use the SIP approach to identify the active micro-organisms in a triclosan-degrading consortium derived from activated sludge exposed to 13C-labelled triclosan. Genomic fingerprinting analyses of resolved 13C-labelled DNA allowed the design of a specific fluorescence in situ hybridization (FISH) probe for a putative triclosan-degrading phylotype, which was used in combination with microautoradiography (MAR) to verify the physiology and abundance of these micro-organisms in the enrichment.
Methods
Reagents and media.
Triclosan (Irgasan) was purchased from Sigma-Aldrich with a purity of >97 %. 13C12-Labelled triclosan (isotope purity >98 %) was purchased from Wellington Laboratories and was dissolved in methanol. [Dichlorophenyl-U-14C]-labelled triclosan (specific activity 5.43 MBq mg−1) was donated by Ciba. Stock solutions of both radiolabelled and unlabelled triclosan were prepared in acetone. As a standard procedure, substrate solutions were allowed to dry at room temperature prior to the addition of specified media or culture. For all experiments, nitrate mineral salts medium (NMS) was used as carbon-free medium (Whittenbury et al., 1970). All other reagents used for SIP and MAR were commercial products of highest grade (Chen et al., 2011; Kristiansen et al., 2011a; Neufeld et al., 2007).
Enrichment culture and growth conditions.
Activated sludge was taken from the aeration tanks from C/N/P-removing Aalborg West WWTP (Aalborg, Denmark) and was used as source of inoculum for enrichment of triclosan-degrading organisms. The initial incubation has been described in a preliminary report (Chen et al., 2011). Briefly, an activated sludge sample was spiked with 2 mg triclosan l−1 and incubated in the dark under aerobic conditions at 22–25 °C on a rotary table (150 r.p.m.). Following the initial incubation and every 9 days thereafter, the enrichment culture was transferred [10 % (v/v)] to fresh NMS medium containing 2 mg triclosan l−1. The enriched culture had a maximum cell density of 6×108 cells ml−1 and was maintained for 4 months before conducting the SIP and MAR incubations.
Analytical methods
Liquid–liquid extraction.
Samples (5 ml) from the experiments were extracted by addition of 2 ml toluene and 100 µl internal standard solution (1000 ng musk xylene D15 ml−1) and were vigorously stirred for 5 min. The organic phase was extracted and the residual water was removed by freezing the samples overnight at −20 °C. These organic extracts were then concentrated to 1 ml with a nitrogen flow condensor at 55 °C.
Instrumental analysis.
Triclosan extracts were finally analysed by gas chromatography with mass spectrometric detection (GC-MS, Thermo-Trace-MS and Trace GC) equipped with a splitless injector and A200S autosampler. Samples (1 µl) were injected into the injector in splitless (1.5 min) mode held at 240 °C. The GC separation was performed with an Rxi-5Sil MS column (Restek): length, 10 m; ID, 0.18 mm; film, 0.18 µm; and a temperature programme of 90 °C (hold 1 min) ramped at 50 °C min−1 to 135 °C and then at 10 °C min−1 to 220 °C. Finally, the baking temperature was reached by ramping the column at 40 °C min−1 to 260 °C which was held for 6 min. Helium (5.0) was used as carrier gas with a flow rate of 1.3 ml min−1. The transfer line of the mass spectrometer (Trace MS, Thermo Finnigan) was held at 250 °C. The ion source was operated at 160 °C. The mass spectrometer was operated in selected ion mode (SIM) utilizing 31–61 ms dwell time. The detector of the mass spectrometer was operated at 450 V. The recovery rate of triclosan was 88±11 % (sd) and limit of quantification was 3 ng g−1, as reported by Bester (2003).
SIP.
A total of 5 ml (approx. 100 mg dry matter) of the enriched culture was transferred to 60 ml serum bottles and incubated with 2 mg 13C-labelled triclosan l−1 for 3 days. Parallel incubations were also prepared with unlabelled substrate and used as controls for verification of DNA-SIP labelling and triclosan degradation. The bottles were crimp-sealed with rubber stoppers and incubated in the dark at 24 °C on a rotary table (150 r.p.m.) for 3 days. Subsequently, total DNA was extracted using the FastDNA SPIN kit for Soil (MP Biomedicals) according to the manufacturer’s instructions. The DNA concentration was measured on a NanoDrop 2000 spectrophotometer (Thermo Fisher Scientific). All incubations were carried out as biological duplicates.
Isolation and fingerprinting of 13C-labelled DNA.
Caesium chloride (CsCl) gradient fractionation, DNA precipitation and DNA quantification were set up as described previously (Neufeld et al., 2007). Briefly, 2 µg DNA from each sample (two control samples and two 13C-labelled samples) was added to the gradient buffer and mixed with CsCl to a final density of 1.725 g ml−1. These solutions were added to 5.1 ml polyallomer Quik-seal centrifuge tubes (Beckman Coulter) and ultracentrifuged at 133 000 gav for 72 h at 20 °C in a Sorvall TH-641 swing-out rotor (Kendro). Immediately after centrifugation, the density gradients were fractionated into 12 volumes of approximately 400 µl. The buoyant density of each fraction was determined by measuring 5 µl from each sample on a refractometer (AR200, Reichert). DNA from each fraction was precipitated with polyethylene glycol and glycogen as described elsewhere (Neufeld et al., 2007), and followed by resuspension in nuclease-free water. DNA was quantified using a NanoDrop 2000 spectrophotometer.
The shift in community between the control and the labelled fraction was visualized by molecular profiling using denaturing gradient gel electrophoresis (DGGE) and PCR. DGGE was performed as described in detail elsewhere (Kristiansen et al., 2011a). From DGGE results, distinct DNA bands from the labelled heavy fractions (buoyant density 1.83 and 1.79 g ml−1) were chosen for subsequent sequencing (Fig. S1, available with the online version of this paper). Furthermore, the 12C and 13C-labelled DNA fractions were used as template for PCR with the 16S rRNA gene-targeted primers 26F/1492R (approx. 1450 bp product) (Lane, 1991). PCR conditions are described elsewhere (Kristiansen et al., 2011b). A 16S rRNA clone library was prepared from the high density fractions (1.76–1.80 g ml−1) of the SIP incubation with the 13C-labelled triclosan. The clone library preparation and the phylogenetic analysis were performed as described by Kristiansen et al. (2011a) except that the alignment and phylogenetic tree construction were done using mega 5 (Tamura et al., 2011). Screening of the clone sequences with Bellerophon v3 (DeSantis et al., 2006) did not identify any putative chimeras. Sequences represented in the phylogenetic tree were named triclosan-degrading culture clones and deposited in the GenBank database under accession numbers JX099503–JX099536.
FISH probe design.
The 16S rRNA gene sequences from the clone library were used to design an oligonucleotide probe (Meth1138) (Table 1) using the probe design tool in the arb software package (Ludwig et al., 2004); the probe was subsequently confirmed for specificity using the check probe programme in the Ribosomal Database Project (Maidak et al., 2000). Optimum hybridization stringency for the probe was determined by performing formamide dissociation series on biomass from the enrichment culture and activated sludge from Aalborg West WWTP with 10 % formamide (v/v) increments across a range of 0–60 % (v/v). Prior to FISH, samples were homogenized and fixed with 4 % (w/v) paraformaldehyde, as described previously (Nielsen, 2009). The group-specific probe Meth1138 was labelled with sulfoindocyanine dyes (Cy3). FISH analysis was performed by using the general bacterial probe mixture EUBmix labelled with 5(6)-carboxyfluorescein-N-hydroxysuccinimide ester (FLUOS) and more specific probes labelled with Cy3 (Table 1). The FISH procedure was carried out as described previously (Nielsen, 2009). An epifluorescence microscope (Axioscope 2, Carl Zeiss) was used in all FISH analyses. Bacterial abundance was quantified by measuring the ratio of the area fluorescing with a probe (Cy3 labelled) to the area fluorescing with EUBmix probe (FLUOS labelled) on the same microscopic field. For each enumeration, at least 20 images were taken from two separate hybridizations and analysed using ImageJ software (Collins, 2007). FISH analyses were also conducted on samples taken from seven Danish WWTPs: Bjergmarken, Aalborg East, Egå, Ejbymølle, Hjørring, Skive and Aalborg West. These plants represent stable and well-functioning C/N/P-removing treatment plants with different configurations and influent wastewater composition.
Microautoradiography.
Microautoradiography experiments in enrichment culture in combination with FISH (MAR-FISH) were performed as described previously (Nielsen & Nielsen, 2005). Briefly, 5 ml of the enriched culture was transferred to 9 ml serum bottles and incubated with 10 µCi 14C-labelled triclosan (3.7×107 Bq) and unlabelled triclosan to a final concentration of 2 mg l−1 under aerobic conditions for 1 day on a rotary table (labelled and unlabelled triclosan was added at time 0). As a control for chemography, a sample from the enriched culture was pasteurized at 70 °C for 10 min prior to MAR incubation and run in parallel. MAR incubations were terminated by fixing samples with 4 % (w/v) paraformaldehyde. The samples were then washed, homogenized and immobilized on gelatin-coated coverslips as described elsewhere (Nielsen & Nielsen, 2005). Finally, the samples were subjected to FISH. After the FISH procedure, the samples were coated with liquid film emulsion (Kodak) and exposed in the dark for 3–6 days before being developed and microscopically examined. Production of 14C-labelled CO2 was monitored in MAR-incubated culture by measuring the percentage accumulation of precipitated radioactivity using a liquid scintillation counter (Packard 1600 TR; Packard) as follows. Samples (1 ml) from the headspace gas were withdrawn using a syringe and mixed with 1 ml 0.1 M NaOH solution in a gas-tight sealed serum bottle. At the same time, 0.1 ml aliquots were withdrawn from the culture and directly transferred to 3 ml scintillation liquid (Ultima Gold XR; Packard) to measure the total radioactivity of the culture. All incubations were carried out as biological duplicates.
Results
Biodegradation of triclosan in enrichment culture
After spiking the enrichment culture with 2 mg triclosan l−1, the concentration of triclosan was reduced below the limit of quantification (3 ng l−1) within 90 h, whereas the triclosan concentration remained nearly constant in a pasteurized control spiked with 1 mg triclosan l−1 (Fig. 1). This indicates that the removal of triclosan was predominantly due to biological activity, which agrees with the literature (Bester, 2005; Singer et al., 2002). Degradation of 2 mg triclosan l−1 followed first-order kinetics with a removal rate and half-life of 0.0431 h−1 and 16 h, respectively. This was approximately five times faster than that calculated from degradation analyses where the same concentration of triclosan was spiked directly into activated sludge from the Aalborg West WWTP (Chen et al., 2011). When radiolabelled triclosan was spiked into the enrichment culture, the subsequent liquid scintillation counts showed that this microbial community is able to mineralize triclosan, with approximately 13 % of the added radioactivity detected in the headspace after 3 days of incubation (Fig. 1). The linear progression of 14C-labelled CO2 in the headspace indicates that triclosan degradation started immediately after its addition without a lag phase (Fig. 1). Meanwhile, no further accumulation of 14C-labelled CO2 was observed when the triclosan concentration was reduced below the detection limit.
Biodegradation of triclosan in the enrichment culture under aerobic conditions. The time-course of triclosan degradation (2 mg l−1) in pasteurized (•) and active (○) enriched culture is shown. Cumulative production of 14C-labelled CO2 (□) from the mineralization of 14C-labelled triclosan is shown on the secondary y-axis. Error bars for triclosan measurements indicate the 10 % stated uncertainty from the method development by Bester (2003).
Detection and phylogenetic analysis of 13C-labelled bacterial 16S rRNA gene sequence
Following incubation with 13C-labelled triclosan, total DNA was extracted and centrifuged in a CsCl density gradient to separate labelled from non-labelled DNA. This resulted in a linear isopycnic gradient from 1.83 to 1.57 g ml−1 (Fig. 2). Although the buoyant densities in our SIP fractionation were relatively broad, a clear shift towards a heavier density of the quantified DNA was observed from both duplicate samples incubated with unlabelled triclosan relative to the 12C-labelled control. This shift was also apparent from the band intensity of the PCR product after 25 cycles of amplification of the 16S rRNA genes (Fig. 2). The shift in density was further evaluated by DGGE, revealing a shift in banding patterns in the heavy fractions (density 1.80–1.76 g ml−1) of the 13C-labelled triclosan incubation compared with the unlabelled control (Fig. S1). Distinct bands (1, 3, 6, 7, 8 and 9 on Fig. S1) were identified as Methylobacillus sp. Iva (GU937479), while the remaining bands (2, 4 and 5) were unclassified.
Quantitative profiles of DNA in CsCl SIP gradients. The distribution of DNA in comparative SIP gradients of DNA from microcosms (n = 2) amended with 12C- or 13C-labelled triclosan is shown in (a). Open symbols show the results of the biological duplicate. (b) Band intensity of PCR products targeting the 16S rRNA genes (~1.5 kb) amplified from the corresponding gradient fractions and visualized after agarose gel electrophoresis.
Thirty four clones of PCR-amplified 16S rRNA genes were sequenced from the 13C-enriched DNA fractions. Most of the clones (31 of 34 sequences) affiliated with the genus Methylobacillus (Fig. 3). The obtained sequences had less than 95 % identity to the other previously described Methylobacillus species. Three clone sequences were related to the genus Stenotrophomonas within the Gammaproteobacteria (Fig. 3) with strong bootstrap support and less than 95 % identity to other Stenotrophomonas sequences.
Phylogenetic affiliation of the 16S rRNA gene sequences obtained from the 13C-enriched SIP fractions. GenBank accession nos or the number of clone sequences obtained are indicated in parentheses. The tree was constructed using the maximum-likelihood algorithm with branching confidence values from 1000 replicates. Bootstrap values ≥75 % and ≥90 % are indicated by empty and filled circles, respectively. Bar, 5 % sequence divergence; the outgroup was made from 10 randomly chosen Chloroflexi gene sequences.
Identification of triclosan-utilizing bacteria
To verify that members of the genus Methylobacillus were utilizing triclosan in the enrichment culture, a specific FISH probe (Meth1138) targeting most members of the genera was designed. The hybridization stringency of the probe was optimized on biomass from the culture and from activated sludge samples and determined to be 25 % (v/v) formamide. The probe was used to quantify the relative abundance of Methylobacillus in the enriched culture as well as in activated sludge and was calculated to range between 2 and 4 % and 0.5 and 1 % of the total detected cells, respectively. No further enrichment was detected during SIP or MAR incubations. Dense silver grain patches covering the Meth1138-hybridized cells indicated an active utilization of 14C-labelled triclosan by Methylobacillus (Fig. 4). A low background in the MAR visualizations and lack of MAR-positive cells in the pasteurized control indicated a low absorbance and chemography of 14C-labelled triclosan to the sample. In the enrichment culture, approximately 25 % of the Meth1138-positive cells were MAR-positive, but other betaproteobacterial cells (positive with the BET42a probe) were also MAR-positive (Fig. 4). These cells were found to constitute 2–3 % of the total number of cells detected by EUBmix and gave similar silver grain density to the MAR-positive Methylobacillus.
FISH and microautoradiography images of triclosan-utilizing cells. FISH images of the triclosan-fed enrichment culture after hybridization with (a) the universal bacterial probe EUBmix (green) and probe Meth1138 (red), and (b) probe BET42A (green) and probe GAM42a (red). Cells appearing yellow hybridized with both probes. Representative MAR-FISH images of triclosan-utilizing bacteria present in the enrichment culture incubated for 1 day at 1 mg 14C-labelled triclosan l−1 and hybridized with BET42A probe (red) (c, e) or Meth1138 (red) and EUBmix probe (green) (d, f–i). Silver grains surrounding bacterial cells indicate active cellular incorporation of the 14C-labelled triclosan (white arrows), while black arrows indicate MAR-negative cells. Bars, 10 µm.
The oligonucleotide probe was applied to assess the abundance of FISH-detectable Methylobacillus bacteria in seven Danish full-scale wastewater treatment plants. With a detection limit of 0.25 % of the biovolume, estimated by the use of nonsense probe NONEUB (Wallner et al., 1993), the survey revealed a highly variable presence of bacteria affiliated with Methylobacillus; some plants showed a complete absence or around the limit of quantification (P<0.1; Aalborg East, Egå, Hjørring, Skive WWTPs) while others showed relatively high abundance (0.5–2 %, P<0.05, Bjergmarken, Ejbymølle, Aalborg West WWTPs). The probe hybridized with small, rod-shaped cells (Fig. 4) that had similar morphology in samples from the enrichment culture and all the activated sludge WWTPs (Fig. 4).
MAR-FISH was also attempted with biomass from a full-scale plant to confirm that these organisms are involved in triclosan removal in these systems. However, due to the presence of very few MAR-positive cells (enumerated to be around 2 % of the total number of cells detected by EUBmix, corresponding to approximately 8×106 cells ml−1) combined with low fluorescence intensities we were not able to assess with confidence the MAR-FISH signals. Design of specific FISH probes targeting the three Stenotrophomonas sequences identified by SIP failed to detect target cells and previously published probes for this genus had one mismatch to the sequences obtained from the clone library. However, due to the relatively low abundance of Gammaproteobacteria in the enrichment culture (Fig. 4) [<1 % positive with the GAM42a probe compared with ~95 % of Betaproteobacteria (BET42a) relative to the total FISH positive cells detected by EUBmix] and the observation that all MAR-positive cells were also Betaproteobacteria-positive (Fig. 4), no further attempts were taken to verify if members of the Stenotrophomonas were taking up 14C-labelled triclosan.
Discussion
Although more than 60 % of total removal of triclosan is attributed to the biodegradation processes in activated sludge treatment (Bester 2003, 2005) very little is known about the micro-organisms involved. Previous studies have shown the ability of a few isolates to degrade triclosan (Gangadharan Puthiya Veetil et al., 2012; Hay et al., 2001; Kim et al., 2011; Meade et al., 2001). However, as these studies rely on the use of culture-dependent methods and do not necessarily reflect the identity of the active members involved in the biodegradation of triclosan in situ, the focus of this study was to apply SIP to identify bacteria capable of utilizing triclosan in an enrichment culture. Attempts to apply the SIP approach directly on activated sludge were not successful, most likely because of the low numbers of bacteria involved in the degradation of triclosan as observed in the MAR-FISH results. Apparently, with the amount of 13C-labelled triclosan added and the sequencing approach used, we were unable to reach sufficient density shift for the labelled DNA during SIP. So, in order to identify triclosan degraders, an enrichment step was introduced. This approach is biased to enrich for triclosan degraders with a low affinity for the substrate, and discriminates against cells with a high substrate affinity.
The enrichment culture, originally started from activated sludge, was fed on regular additions of triclosan and was able to degrade 2 mg triclosan l−1 with a half-life of 16 h compared with 90 h in the original activated sludge sample. The relatively stable and high removal rate of triclosan and consecutive development of 14C-labelled CO2 combined with a lack of lag phase in the degradation experiments suggest that the enriched bacterial community has readily adapted to triclosan as a carbon source. The consumption of 13C-labelled triclosan resulted in a sufficient amount of heavy-labelled DNA, and a shift in the average density of total DNA compared with the unlabelled (12C) control. Generally, to ensure sufficient DNA labelling in SIP experiments, a few doubling times with the labelled substrate is required. This potentially raises concern regarding cross-feeding of labelled carbon. However, we applied a relatively short incubation period and low concentration of the applied 13C-labelled triclosan to minimize the risk of cross-feeding. The predominant enrichment of a single bacterial clade, the lack of by-products identified, and the confirmation by MAR-FISH with reduced incubation time and tracer, support that the identified Methylobacillus are the primary consumers of triclosan in the enrichment culture. The methodological approach of applying SIP with MAR-FISH is a powerful combination that validates the SIP findings and ensures correct interpretation of even organisms with low abundance. The MAR approach requires less uptake of tracer compared with SIP, and is therefore more sensitive and can be used to test uptake of substrate in natural systems under in situ conditions. However, we were not able to conclusively verify that Methylobacillus was the main triclosan consumer using in situ concentrations in the indigenous activated sludge sample due to very low numbers of MAR-positive triclosan degraders.
The finding of a few Stenotrophomonas-related clones in the 13C-labelled DNA clone library could indicate a broader diversity of triclosan degraders or the presence of multiple degradation steps catalysed by different micro-organisms. Although members of Stenotrophomonas have previously been shown to be involved in the degradation of environmental pollutants with aromatic structures such as p-nitrophenol (Liu et al., 2007), nonylphenol (Soares et al., 2003) and benzene (Lee et al., 2002), their involvement in degradation of triclosan was not confirmed by the MAR-FISH approach. Another betaproteobacterial group was found to be present in similar numbers to Methylobacillus and with similar triclosan degradation activity; however, these cells were not identified by the SIP approach. This could be due to insufficient density shift in the SIP fractionation.
Other studies have shown that less than 1 % of the triclosan added to activated sludge is actually transformed into triclosan-methyl, and that the increase of triclosan-methyl corresponded to the decrease of the parent compound (Chen et al., 2011). We attempted to find and identify triclosan degradation by-products from the enrichment culture by GC-MS and revealed the presence of 2,4-dichlorophenol but this was below the limit of quantification. The lack of accumulated by-products and development of labelled CO2 in the head space during incubation with 14C-triclosan indicates that the added triclosan was fully mineralized. Alternatively, the findings could suggest the presence of a more metabolically diverse community of triclosan degraders in activated sludge, but these would typically be present in small numbers and therefore difficult to identify.
Methylobacillus belongs to methylotrophs, which is a phenotypically defined group capable of using one-carbon compounds as the sole source of energy and carbon (Hanson & Hanson, 1996). However, it has been shown that several methylotrophs that are within the genus Methylobacillus can degrade organic compounds through co-metabolism, such as the pesticide carbonfuran and choline (Hanson & Hanson, 1996), or through direct metabolism, such as microcystin (Hu et al., 2009). Other methylotrophs are known for their ability to participate in the co-metabolic degradation of various environmental pollutants, including trichloroethylene, phenol and different aromatic compounds (Chongcharoen et al., 2005; Koh et al., 1993; Tsuji et al., 1990). Metabolic pathway analyses have shown that Methylobacillus contains unique clusters of genes encoding the degradation of chlorocatechol, a major intermediate product in the biodegradation of chloroaromatic compounds (Caspi et al., 2012; Spokes & Walker, 1974). Little information is available regarding the biodegradation products of triclosan, although catechol and 3,5-dichlorocatechol were detected when triclosan was degraded by pure cultures of Pseudomonas-like strains (Gangadharan Puthiya Veetil et al., 2012) and Sphingomonas sp. PH-07 incubated with diphenyl ether (Kim et al., 2011). This information supports the notion that members of the genus Methylobacillus may play a role in triclosan degradation in the enriched culture and in WWTPs. To our knowledge, organisms within this group have not previously been linked to triclosan degradation. The FISH surveys in the seven Danish WWTPs show that Methylobacillus are indeed present in activated sludge although they are more abundant than can be ascribed to degradation of micro-pollutants such as triclosan, and the abundance thus indicates that they are involved not only in degrading aromatic micro-pollutants but likely also in other processes as well.
In conclusion, SIP combined with MAR-FISH was used here to identify the active community responsible for the degradation of triclosan within an enrichment culture originating from activated sludge. The findings show the ability of members of the genus Methylobacillus to utilize triclosan. Identifying the specific organisms involved in triclosan degradation provides valuable information that may lead to possible strategies to enhance micro-pollutant removal.
Acknowledgements
This study was financially supported by the Danish Research Council (FTP) through grant no. 09-065064. We thank Simon McIlroy (Aalborg University) for helpful comments on the manuscript and Ciba for providing the 14C-labelled triclosan.