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Aerobic biodegradation of trichloroethylene and phenol co-contaminants in groundwater by a bacterial community using hydrogen peroxide as the sole oxygen source a
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Hui Li , Shi-yang Zhang , Xiao-li Wang , Jie Yang , Ji-dong Gu , Rui-li Zhu , Ping Wang , a
Kuang-fei Lin & Yong-di Liu
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State Environmental Protection Key Laboratory of Environmental Risk Assessment and Control on Chemical Process, State Key Laboratory of Bioreactor Engineering, School of Resources and Environmental Engineering, East China University of Science and Technology, Shanghai 200237, People's Republic of China b
School of Biological Engineering, East China University of Science and Technology, Shanghai 200237, People's Republic of China c
Research Institute of Wastes and Soil Remediation, Shanghai Academy of Environmental Sciences, Shanghai 200233, People's Republic of China d
School of Biological Sciences, Swire Institute of Marine Science, The University of Hong Kong, Pokfulam Road, Hong Kong SAR, People's Republic of China Accepted author version posted online: 15 Sep 2014.Published online: 22 Sep 2014.
To cite this article: Hui Li, Shi-yang Zhang, Xiao-li Wang, Jie Yang, Ji-dong Gu, Rui-li Zhu, Ping Wang, Kuang-fei Lin & Yong-di Liu (2014): Aerobic biodegradation of trichloroethylene and phenol co-contaminants in groundwater by a bacterial community using hydrogen peroxide as the sole oxygen source, Environmental Technology, DOI: 10.1080/09593330.2014.957730 To link to this article: http://dx.doi.org/10.1080/09593330.2014.957730
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Environmental Technology, 2014 http://dx.doi.org/10.1080/09593330.2014.957730
Aerobic biodegradation of trichloroethylene and phenol co-contaminants in groundwater by a bacterial community using hydrogen peroxide as the sole oxygen source Hui Lia,† , Shi-yang Zhanga,† , Xiao-li Wanga,b , Jie Yangc , Ji-dong Gud , Rui-li Zhua , Ping Wanga,b , Kuang-fei Lina and Yong-di Liua∗
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a State Environmental Protection Key Laboratory of Environmental Risk Assessment and Control on Chemical Process, State Key Laboratory of Bioreactor Engineering, School of Resources and Environmental Engineering, East China University of Science and Technology, Shanghai 200237, People’s Republic of China; b School of Biological Engineering, East China University of Science and Technology, Shanghai 200237, People’s Republic of China; c Research Institute of Wastes and Soil Remediation, Shanghai Academy of Environmental Sciences, Shanghai 200233, People’s Republic of China; d School of Biological Sciences, Swire Institute of Marine Science, The University of Hong Kong, Pokfulam Road, Hong Kong SAR, People’s Republic of China
(Received 26 February 2014; final version received 1 August 2014 ) Trichloroethylene (TCE) and phenol were often found together as co-contaminants in the groundwater of industrial contaminated sites. An effective method to remove TCE was aerobic biodegradation by co-metabolism using phenol as growth substrates. However, the aerobic biodegradation process was easily limited by low concentration of dissolved oxygen (DO) in groundwater, and DO was improved by air blast technique with difficulty. This study enriched a bacterial community using hydrogen peroxide (H2 O2 ) as the sole oxygen source to aerobically degrade TCE by co-metabolism with phenol in groundwater. The enriched cultures were acclimatized to 2–8 mM H2 O2 which induced catalase, superoxide dismutase and peroxidase to decompose H2 O2 to release O2 and reduce the toxicity. The bacterial community could degrade 120 mg/L TCE within 12 days by using 8 mM H2 O2 as the optimum concentration, and the TCE degradation efficiency reached up to 80.6%. 16S rRNA gene cloning and sequencing showed that Bordetella, Stenotrophomonas sp., Sinorhizobium sp., Variovorax sp. and Sphingobium sp. were the dominant species in the enrichments, which were clustered in three phyla: Alphaproteobacteria, Betaproteobacteria and Gammaproteobacteria. Polymerase chain reaction detection proved that phenol hydroxylase (Lph) gene was involved in the co-metabolic degradation of phenol and TCE, which indicated that hydroxylase might catalyse the epoxidation of TCE to form the unstable molecule TCE-epoxide. The findings are significant for understanding the mechanism of biodegradation of TCE and phenol co-contamination and helpful for the potential applications of an aerobic bioremediation in situ the contaminated sites. Keywords: chlorinated hydrocarbon; phenol; aerobic biodegradation; hydrogen peroxide; bacterial community
1. Introduction Trichloroethylene (TCE) is a common chlorinated solvent used in industrial processes, and is a widespread contaminant present in soil and groundwater because of improper disposal practices for spent solvents in the past.[1] Meanwhile, amounts of phenols and their close relatives were also disposed from chemical industries, which resulted in the co-contaminations of TCE and phenols distributed in many chemical industrial polluted sites. Biological treatments are considered to be the most effective for the remediation of TCE and phenol co-contaminants because they allow efficient and complete removal from the environment. The TCE biodegradation pathway involves either reductive dechlorination under anaerobic conditions or co-metabolism under aerobic conditions. However, more toxic chemicals such as dichloroethylene and vinyl chloride are generated during the anaerobic process.[2] On
*Corresponding author. Email: [email protected] † Shi-yang Zhang and Hui Li contributed equally to this article. © 2014 Taylor & Francis
the other hand, TCE is partially mineralized to non-toxic materials such as water, carbon dioxide and chlorine ions under aerobic conditions.[3] Although bacterial growth cannot be sustained using TCE as the carbon and energy source, some bacteria utilizing various growth substrates, including methane, phenol, toluene and ammonia, are capable of utilizing TCE as the secondary substrate via cometabolism.[4–6] Groundwater contaminated with TCE has been treated in situ via co-metabolic biodegradation by injecting phenol or toluene.[7] Furthermore, the TCE co-metabolic pathway involving the Lph gene is well established. Lph catalyses the conversion of phenol to catechol and simultaneously co-metabolizes TCE. Chen et al. reported the aerobic degradation process of TCE by phenol-oxidizing bacteria, Pseudomonas putida, which showed a degradation rate of 0.015 mg TCE/mg dry weight/h.[8] Moreover, TCE bioremediation studies
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at contaminated sites and in groundwater have been conducted by adding advantageous bacteria, including the common phenol-utilizing bacteria.[9] Because microorganisms need dissolved oxygen (DO) as the electron acceptor during aerobic biodegradation, the absence of DO in the groundwater renders the aerobic degradation of pollutants difficult. The available oxygen in the groundwater environment could be increased by introducing oxygen into the aquifer by sparging with air, pure oxygen or ozone and by the addition of H2 O2 or oxygen-releasing compounds.[10,11] Use of H2 O2 is a cost-effective approach that is attracting increased attention, because it is soluble in water and provides unlimited amounts of oxygen for biodegradation. Mukherjee et al. have shown that the accretion of 0.01% (v/v) H2 O2 as an additional source of oxygen in the culture medium containing Bacillus subtilis DM-04 and Pseudomonas aeruginosa strains remarkably enhanced the biodegradation rate of benzene, toluene and three xylene isomers under oxygen-limited conditions.[12] However, H2 O2 showed an inhibitory or toxic effect on cell growth with a longer lag phase and slightly lower specific growth rate. Numerous studies have revealed that H2 O2 is toxic to some microorganisms at concentrations as low as 0.003%.[13] Despite the apparent contradiction of using a potential biological toxicant to facilitate aerobic biodegradation, subsurface bioremediation by supplying H2 O2 has been continued on the premise that H2 O2 toxicity would not damage the biodegradation process. Fortunately, aerobic microorganisms have adaptive mechanisms to overcome the adverse effects of external exposure to H2 O2 . Aerobic microorganisms possess catalase (CAT), which decomposes H2 O2 ; superoxide dismutase (SOD), which can remove superoxide ions (O2− ) generated enzymatically; and peroxidase (POD), which uses H2 O2 to oxidize organic compounds.[14] CAT and SOD function cooperatively for protecting microorganisms against H2 O2 toxicity, whereas POD plays a secondary role in the defence mechanism. Increased antioxidase-specific activity was considered to be an indicator of adaptation to H2 O2 under elevated H2 O2 conditions. H2 O2 induces CAT in nearly all aerobic organisms such as Bacillus sp. and Serratia marcescens and induces SOD in Lactococcus spp. and Streptococcus thermophilus.[15,16] H2 O2 has been shown to markedly affect the survival of microorganisms; however, relatively little is known regarding the structure of a microbial community grown in the presence of H2 O2 . This study aimed to obtain a bacterial community that degraded TCE and phenol co-contaminants in the presence of phenol and H2 O2 . An aerobic enrichment was isolated from a site that was exposed to long-term chlorinated hydrocarbon and phenol pollution, and then the enrichment was acclimated to high concentrations of H2 O2 that served as an oxygen source. In addition, the antioxidase activities of the bacteria were determined to confirm
their adaptability to H2 O2 . Finally, the dominant species of the bacterial community were determined, and the mechanism of TCE and phenol co-metabolism was clarified when H2 O2 was used as sole oxygen source.
2. Materials and methods 2.1. Enrichment of the microbial community A soil sample was obtained from a TCE and phenol co-contaminated site with a depth of 0–1.5 m (Pudong, Shanghai, China). For the aerobic culture, 5 g of the soil sample was added to a 125-mL serum bottle containing 50 mL of chloridion-free MCl medium (Na2 HPO4 7.0 g, KH2 PO4 3.0 g, NH4 H2 PO4 2.52 g, Ca(NO3 )2 ·4H2 O 0.024 g, MgSO4 ·7H2 O 0.246 g, H2 O 1000 mL, pH 7.0), followed by the addition of 5 mg/L TCE and 80 mg/L phenol as co-metabolic substances. The serum vials were sealed with butyl stoppers, which were crimped with aluminium caps. Autoclaved vials prepared as above were used as controls. The vials were then cultured at 30°C and shaken at 120 rpm. The enriched cultures were transferred 4–5 times and, simultaneously, the concentration of TCE was constantly increased. Subsequently, 20 mL of the previously mentioned well-grown mixed culture was inoculated into the vials as described above, but amended with 80 mL MCl medium. The resultant solution was cooled under a stream of nitrogen in order to remove oxygen, and then 20 mg/L TCE, 80 mg/L phenol and 2 mM H2 O2 were added. The enriched cultures were acclimatized to H2 O2 by gradually increasing the concentrations of H2 O2 (from 2 to 8 mM in 2 months). The effect on TCE biodegradation (concentration, 20–120 mg/L) in the presence of H2 O2 was also assessed in this study. Cell growth was determined against the control by measuring the optical density at 600 nm (OD600 ).
2.2. TCE degradation analysis TCE biodegradation was monitored using gas chromatography (GC). For this, a 1 mL sample was obtained from each vial by using a sterile syringe and mixed with 1 mL n-hexane. After the sample was oscillated and extracted, 1 μL organic phase was analysed using GC by using Agilent A7890 equipped with an electron capture detector. The injector, oven and detector temperatures were set at 240°C, 80°C and 260°C, respectively. Nitrogen (flow rate, 20 mL/min) was used as the carrier gas. Samples were injected at a split rate of 5:1, and the retention time of TCE was approximately 5.1 min. The concentration of TCE degradation was corrected for the volatilized amount of TCE by using the following equation: TCEc = TCEM + rvol.TCE × t, where TCEC , TCEM , rvol.TCE and t are corrected TCE concentration (ppm), measured TCE concentration (ppm), TCE volatilization rate (ppm/day) and time in days, respectively.
Environmental Technology 2.3. H2 O2 and DO analyses The titanium oxalate method was used to spectrophotometrically quantify the H2 O2 levels.[17] The absorbance at 400 nm was measured using a UV spectrophotometer (UV 1800; SHIMADZU). The assay was used for H2 O2 concentration ranging from 0 to 12 mM. DO of the samples was detected using a Clark-type electrode (YSI 550-12 Oxygen Meter; Yellow Springs). A three ml sample was injected to the 10 mL-reaction vessel and the probe was inserted into the liquid. The reaction vessel was sealed with Teflon film. A wait of 10 min ensued thereafter by setting the electronic timer, to get a stable reading and write down the measurement of conductivity.
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2.4.
Key enzyme activity assays
The protein concentration of the bacterial suspension was determined using the Coomassie blue method.[18] Coomassie Brilliant Blue G-250 and bovine serum albumin were used for the generation of standard curves. The related enzyme activities were measured by preparing crude extracts. For this, 10 mL culture was harvested by centrifugation (4000 g, 10 min, 4°C), and the supernatant was removed. Subsequently, the cells and 10 mL phosphate buffer solution were mixed using a vortex mixer for a few seconds. The washing process was repeated three times. Next, the cell suspensions were sonicated for 10 min at 45-s bursts at an output level of 6 by using a cup horn attachment, and then the centrifuge tubes were maintained below 4°C. Cell debris was removed by centrifugation at 12,000 g; the supernatant was used for enzyme assays. CAT activity was assayed as the rate of decrease of the absorbance resulting from the disappearance of 0.01 mol/L H2 O2 solution (0.05 mol/L phosphate buffer; pH 7.0; 25°C) at 240 nm. One unit of enzyme activity is the amount of CAT that degrades 1 mmol of H2 O2 per minute. ® The SOD activity was assayed using an SOD assay-WST kit (Dojindo, Japan). The enzyme and WST-1 working solutions and crude extracts were prepared according to the manufacturer’s instruction. After all the solutions were added to 96-well microplates as described by the manufacturer, the microplates were incubated at 37°C for 20 min. Subsequently, the absorbance was measured at 450 nm by using a microplate reader (Sunrise, Tecan, Austria). One unit of SOD activity was defined as the amount of enzyme in a sample solution causing 50% inhibition (IC50 ) of the rate of tetrazolium reduction. POD activity was assayed using the method in which the rate of formation of coloured product resulting from the oxidation of guaiacol is measured at 460 nm. The buffer contained 0.2 mol/L phosphate buffer (pH 7.8). The reaction mixture contained 0.1 mol/L phosphate buffer (pH 6.0), 4 mmol/L H2 O2 and 0.6 mg/mL guaiacol. For the assay, 1 mL of the enzyme extract was mixed with 3 mL of the reaction mixture. One unit of enzyme activity was defined as change rate in the absorbance of 45 s.
2.5.
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DNA extraction, cloning and sequencing
After domestication, 10 mL of aerobic medium exposed to H2 O2 was filtered through a sterile 0.22-μm membrane. DNA from the cells was extracted using a bead beating technique (Biospec, Shanghai, China) by using a Fast DNA spin kit for DNA extraction (Abigen, Beijing, China). Polymerase chain reaction (PCR) amplifications were performed using a 25-μL reaction mixture that contained 0.5 μL of template, 0.5 μL of each primer (10 mmol/L; 27F: 5 -AGAGTTTGATCCTGGCTCAG-3 ; 1492R: 5 TACGGTTACCTTGTTACGACTT-3 ), 12.5 μL Taq mix and 11 μL of ddH2 O. The PCR was carried out using the following cycling conditions: 1 cycle of initial denaturation at 94°C for 5 min, denaturation at 94°C for 1 min, annealing at 52°C for 30 s, extension at 72°C for 1 min; 30 cycles of denaturation at 94°C for 1 min, annealing at 52°C for 30 s, extension at 72°C for 1 min; 1 cycle of re-extension at 72°C for 10 min and a final storage at 4°C. The PCR was performed using a Mastercyle gradient (Eppendorf, US). The PCR products were purified using the DNA purification kit (V-gene; Shanghai, China) and then ligated into the pMD 19-T vector system according to the manufacturer’s instructions (TaKaRa, Dalian, China). The ligated plasmids were transformed into Escherichia coli DH5α. The insertion of 16S rRNA gene was screened and retrieved by PCR amplification by using the primer set of M13-47 (5 -CGCCAGGGTTTTCCCAGTCACGAC-3 ) and RVM (5 -GAGCGGATAACAATTTCACACAGG-3 ). The positive clones were sequenced (BGI, Beijing, China). After the sequences were edited and checked manually, the identity of the typical 16S rRNA gene sequences was checked against published sequences in the GenBank database of National Center for Biotechnology Information (http://www.ncbi.nlm.nih.gov) by using the BLAST software. Phylogenetic trees were constructed using Clustal X (1.8) and Molecular Evolutionary Genetics Analysis (MEGA, version 5.05). Robustness of individual branches was estimated by bootstrapping based on 1000 replications.
2.6.
Analysis of the phenol hydroxylase gene involved in TCE co-metabolism
The isolated DNA was used for the amplification of the phenol hydroxylase (Lph) gene that is known to be involved in the phenol-TCE co-metabolic pathway. The gene was amplified using the following primers sets: Lphf (5 -CGCCAGAACCATTTATCGATC-3 ) and Lphr (5 -AGGCATCAAGATCACCGACTG-3 ).
2.7. Nucleotide sequence accession numbers The GenBank accession numbers for bacterial 16S rRNA gene sequences are KF926834 and KF926836–KF926839 and that for the functional gene is KJ004664.
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H. Li et al. the degradation efficiency increased up to 80.6% with the degradation rate of 8.06 ppm/day when the medium were supplied with 8 mmol/L H2 O2 . The bacterial community showed increased tolerance to TCE from 20 to 120 mg/L and the degradation capacity of TCE also increased, which might have contributed to treat the in situ TCEcontaminated groundwater. Meanwhile, the enrichment was able to degrade phenol at H2 O2 concentrations from 2 to 8 mM (Figure 2). Phenol was degraded after 2–4 days and most of phenol biodegradation was achieved when H2 O2 was added to the medium. Finally, the enrichment almost completely degraded 220 mg/L phenol when 8 mM H2 O2 served as the oxygen source. Previous study documented that only the lab-scale biodegradation of TCE occurred under sufficient oxygen environment. Shukla et al. [19] isolated a methanotrophic community from an uncontaminated tropical soil to degrade TCE. The community degraded 50% TCE in 8 days from an initial concentration of 100 ppm.[19] Tartakovsky et al. [20] showed that the performance of a TCE-degrading microbial consortium was improved by using H2 O2 as an oxygen source in a biofilm reactor. The consortium degraded 18 mg/L TCE per day with supplementation of 0.7 g H2 O2 per day).[20] Compared to the previous study, our findings showed a higher TCE biodegradability as well as H2 O2 tolerance in the phenolloaded media. Few studies have reported the degradation of higher concentrations of TCE by phenol-utilizing bacteria in the presence of H2 O2 .
3.2. Figure 1. Variation of TCE concentration and OD600 during the bacterial community acclimating 2–8 mM H2 O2 . The numbers on the top represent the concentration of H2 O2 in each period.
Decomposition of H2 O2 to produce DO Figure 3 showed the variation of H2 O2 concentration and DO during the bacterial community acclimation. When 2 mM H2 O2 was introduced into the medium, the concentration of DO reached 6 mg/L, which indicated that
8 mM H2O2
4 mM
2 mM 240
Concentration of Phenol
200 160
CPhenol /mg/L
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3. Results and discussion 3.1. Acclimating bacterial community to degrade TCE and phenol co-contaminants The enrichment was acclimated to degrade the cocontaminants of 120 mg/L TCE and 220 mg/L phenol in the presence of increasing H2 O2 concentrations at 2–8 mM in 2 months (Figures 1 and 2). At low concentrations of H2 O2 (2 mmol/L), the cell grew rapidly after a short lag phase. Cells continued to grow and reached the highest OD600 when 8 mmol/L H2 O2 was added. Nevertheless, OD600 reached 0.1 after the addition of 12 mmol/L H2 O2 (data not shown). The enriched medium possessed the ability to biodegrade 20–120 mg/L TCE after the addition of H2 O2 . Initially, 20 mg/L TCE was almost reduced by 68% within 6 days when 2 mmol/L H2 O2 was added. Finally,
120 80 40 0
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Figure 2. Variation of phenol concentration during the bacterial community acclimating 2–8 mM H2 O2 . The numbers on the top represent the concentration of H2 O2 in each period.
Figure 3. Variation of H2 O2 concentration and DO during the bacterial community acclimation. The numbers on the top represent the concentration of H2 O2 in each period.
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Environmental Technology H2 O2 was decomposed to release the DO. On the eighth day, the concentration of DO was considerably lower than before, indicating that the bacteria could adapt to 2 mmol/L H2 O2 and extract higher oxygen from H2 O2 . Similarly, the observed DO values after treatment with 8 mmol/L H2 O2 increased to 20 mg/L, and then slowly dropped to 5 mg/L finally. H2 O2 was rarely detected in the entire culture process, but H2 O2 decomposition became considerably slow and gas production increased remarkably after the addition of 12 mM H2 O2 (data not shown). The results indicated that DO would not be the limiting factor for TCE co-metabolism. Supplementation of H2 O2 in increasing doses to the microorganisms allowed them to adapt to H2 O2 . The DO concentration decreased in every phase, indicating that microorganisms need to adapt to H2 O2 . The optimized concentration of H2 O2 was 8 mM in view of the tolerance of H2 O2 and the available oxygen production. 3.3.
Tolerance mechanism of the bacterial community to H2 O2
The activities of the three antioxidases, CAT, SOD and POD, of the enriched culture were influenced by various concentrations of H2 O2 (Table 1). The CAT levels in the presence of 2–8 mmol/L H2 O2 continued to increase and were greater than the value after treatment with 12 mmol/L H2 O2 . Similar to CAT, SOD activity was the highest at 8 mmol/L H2 O2 . The SOD activity at 12 mmol/L H2 O2 was slightly lower than that detected after treatment with other concentrations of H2 O2 , similar to that noted with CAT. The POD activity varied between 0.045 and 0.138 U/mg for the four concentrations of H2 O2 . The difference might seem minor, but the highest POD activity was also observed when 8 mmol/L H2 O2 was added. These three enzymes were often used to evaluate the antioxidant property of bacteria and plants.[21,22] Our data suggest that CAT, SOD and POD activities increased after treatment with 2–8 mmol/L H2 O2 , indicating the tolerance and adaption of the bacteria in the enriched medium to H2 O2 . Nevertheless, the activities of these three enzymes were inhibited remarkably after treatment with 12 mmol/L H2 O2 , which indicated that the enriched culture could not withstand the supplementation of 12 mmol/L H2 O2 . In the previous studies, CAT and SOD were also considered as the main antioxidant enzymes in aerobic organisms.[23] Table 1.
The increases in antioxidase activities of the enriched culture after treatment with 2–8 mmol/L H2 O2 might be attributed to a shift in the bacterial community towards members having intrinsically higher antioxidant activity levels, or to the induction of antioxidases by specific members of the microbial population. The bacterial consortium in this study overcame the toxic effects of H2 O2 by inducing antioxidases when grown in the H2 O2 -supplemented environment. 3.4. The bacterial community structure The community structure of the acclimated enrichment was investigated by developing a clone library, and a total of 30 positive clones from bacterial 16S rRNA gene clone libraries were clustered into five operative taxonomic units (OTUs) with 97% similarity by using Muthor 1.32. The rarefaction curve became flat to the right, indicating that a reasonable number of individuals of the community were sampled. The phylogenetic tree elucidated the dominant species in the enrichment distributed in the three major groups: Betaproteobacteria (19/30, 63.33%, 2 OTUs), Alphaproteobacteria (6/30, 20%, 2 OTUs) and Gammaproteobacteria (5/30, 16.67%, 1 OTU; Figure 4). Clh4 showed high similarity (99%) to Bordetella sp. KP22 (AB015607), which was isolated from the polluted soil with TCE contamination and showed strong TCE-degrading ability. Clh2 exhibited 98% similarity to aerobic para-nitrophenol degrader Stenotrophomonas sp. CERAR5 (JF346409), which was found in a TCE-contaminated site. Clh3 exhibited 96% similarity to Sinorhizobium sp. c37 (AB167207), which was found to be responsible for TCE degradation in a phenol-oxidizing culture.[24] Clh1 exhibited 99% similarity to a bacterium clone ccslm224 (AY133088), which was isolated from a TCE-polluted site. Clh5 exhibited 99% similarity to Sphingobium sp. XTT-3 (JN120899) that could degrade phenol. Bordetella sp. Clh4 was the most dominant species in the enrichment, and Bordetella are known to produce SOD and CAT to detoxify superoxide and H2 O2 .[25] Further, some Stenotrophomonas sp. were pollutant-degrading bacteria having high antioxidant activity.[26] Analysis of the dominant species explained why the enrichment showed considerably higher TCEdegrading activities and was adapted to H2 O2 during the acclimation.
The antioxidase activities after treatment with different concentrations of H2 O2 .
CAT per U/mg protein SOD per U/mg protein POD per U/mg protein
2 mmol/L
4 mmol/L
8 mmol/L
12 mmol/L
2.463 ± 0.317 0.492 ± 0.108 0.047 ± 0.011
2.534 ± 0.248 0.522 ± 0.131 0.065 ± 0.014
3.725 ± 0.286 0.916 ± 0.129 0.138 ± 0.012
0.932 ± 0.195 0.239 ± 0.093 0.045 ± 0.012
Results represent mean ± SD for three independent experiments. p < .05.
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H. Li et al. 100
Clh4 Bordetella sp. KP22 (AB015607)
68
Oxalobacteraceae bacterium HTCC315 (AY429715) Burkholderia cepacia strain AW201 (EF373556)
99
Clh1
92
clone ccslm224 (AY133088) Variovorax paradoxus MBIC3839 (AB008000)
100 80
Clh2 Stenotrophomonas sp. CERAR5 (JF346409) Stenotrophomonas maltophilia strain:c309 (AB167179) Stenotrophomonas maltophilia strain PM102 (JQ797560) 100
Clh5 Sphingobium sp. XTT-3 (JN120899)
66 91
clone 290 (DQ158105) Sphingomonas sp. FG03 (EU784670) Clh3
100
Rhizobium sp. c21 (AB167200)
99
Alphaproteobacteria
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97
Gammaproteobacteria
clone GPT1a (AY706440)
96
100
Betaproteobacteria
71
Sinorhizobium sp. c37 (AB167207)
81 99
Sinorhizobium sp. c93 (AB167234)
0.02 Figure 4. Phylogenetic tree of the V6 partial sequence of bacterial 16S rDNA from the bacteria grown in a culture medium supplemented with 8 mmol/L H2 O2 is shown in boldface, and the referred sequences in the EMBL database with the putative divisions are listed to the right. The topology was calculated using the neighbour-joining method. Bootstrap values n = 1000 replicates. More than 50% of the values were reported near the corresponding nodes. The scale bar represents 0.02 nuclear acid substitutions per nucleotide position.
3.5.
Mechanism of TCE and phenol co-metabolism by using H2 O2 as sole oxygen source Whether TCE degradation by the bacterial community grown in a H2 O2 -supplemented environment proceeded via the known pathways [27] was determined by performing PCR analysis of key genotypes by using primer sets for gene sequences encoding Lph, which is involved in the epoxidation of TCE. The PCR products showed the expected size for the Lph gene (684 bp) for the test sample; further, there was no expression found in the negative samples (Figure 5). The sequencing and BLAST analysis revealed that the sequence had 85% similarity to the Lph genes involved in the TCE-degradation pathway. The epoxidation of TCE was the key metabolic reaction, and many published strains were known to effectively degrade TCE via this pathway, such as Ralstonia eutropha, Variovorax sp. and Pseudomonas sp.[28] The present study showed that the Lph gene was actively involved in the TCE co-metabolism under an H2 O2 -supplemented condition. The results suggested that the mechanism of degrading
Figure 5. PCR product analogues of the indicated catabolic genotypes obtained using DNA extracts of bacteria grown in a medium containing 8 mmol/L H2 O2 by using primer set Lph.
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Figure 6. The mechanism of TCE and phenol co-metabolism in the presence of H2 O2 .
TCE was improved by adding H2 O2 (Figure 6). H2 O2 induced antioxidase enzymes of the bacteria to release O2 ; further, the antioxidases relieved the toxicity to the bacteria caused by H2 O2 and O2− . In addition, the bacterial community was stimulated by phenol to produce Lph enzymes which catalysed the reaction for phenol oxidation. Because the Lph enzyme was an iosenzyme, it catalysed the reactions for O2 and TCE forming TCE-epoxide which was a chemically unstable molecule. The TCE-epoxide spontaneously decomposed into readily degradable compounds, and finally they were degraded to CO2 , H2 O and chloride ions. 4.
Conclusions
In this study, a bacterial community was enriched from TCE-contaminated soil to degrade TCE in the presence of H2 O2 . The H2 O2 -tolerant mechanism of the enrichment involved increasing the activities of antioxidant enzymes. The bacterial community included: Bordetella, Stenotrophomonas, Sinorhizobium, Variovorax and Sphingobium. The dominant species belonged to Bordetella and Stenotrophomonas. The bacterial community induced antioxidase enzymes to release O2 to improve the DO in the system. The Lph enzymes catalysed O2 and TCE to form TCE-epoxide which was readily degradable compounds. This study contributed to the potential applications of an aerobic process for in situ bioremediation of TCE-contaminated groundwater. Acknowledgements This work was supported jointly by National Natural Science Foundation of China (51378208, 41273109, 41003031), Specialized Research Fund for the Doctoral Program of Higher Education (20110074130002), Shanghai Rising-Star Program (12QA1400800), Fok Ying Tung Education Foundation (141077), Program for New Century Excellent Talents in University (NCET-13-0797), Innovation Program of Shanghai Municipal Education Commission (14ZZ059), Fundamental Research Funds for the Central Universities (222201313008). We also
would like to thank the anonymous referees for their helpful comments on this paper.
References [1] Yuan B, Li F, Chen Y, Fu ML. Laboratory-scale column study for remediation of TCE-contaminated aquifers using three-section controlled-release potassium permanganate barriers. J Environ Sci. 2013;25(5):971–977. [2] Hu M, Zhang Y, Liu Y, Wang X, Wong PK. Effect of different nutrients on the anaerobic degradation of trichloroethene at optimal temperature. Int Biodeter Biodegr. 2013;85:103– 107. [3] Shukla AK, Upadhyay SN, Dubey SK. Current trends in trichloroethylene biodegradation: a review. Crit Rev Biotechnol. 2012;34(2):1–15. [4] Kocamemi BA, Cecen F. Biological removal of the xenobiotic trichloroethylene (TCE) through cometabolism in nitrifying systems. Bioresource Technol. 2010;101(1):430– 433. [5] Liu J, Amemiya T, Chang Q, Qian Y, Itoh K. Toluene dioxygenase expression correlates with trichloroethylene degradation capacity in Pseudomonas putida F1 cultures. Biodegradation. 2012;23(5):683–691. [6] Powell C, Nogaro G, Agrawal A. Aerobic cometabolic degradation of trichloroethene by methane and ammonia oxidizing microorganisms naturally associated with Carex comosa roots. Biodegradation. 2011;22(3):527–538. [7] Goltz MN, Gandhi RK, Gorelick SM, Hopkins GD, Smith LH, Timmins BH, McCarty PL. Field evaluation of in situ source reduction of trichloroethylene in groundwater using bioenhanced in-well vapor stripping. Environ Sci Technol. 2005;39(22):8963–8970. [8] Chen YM, Lin TF, Huang C, Lin JC. Cometabolic degradation kinetics of TCE and phenol by Pseudomonas putida. Chemosphere. 2008;72(11):1671–1680. [9] Pant P, Pant S. A review: advances in microbial remediation of trichloroethylene (TCE). J Environ Sci. 2010;22(1):116– 126. [10] Hu L, Wu X, Liu Y, Meegoda JN, Gao S. Physical modeling of air flow during air sparging remediation. Environ Sci Technol. 2010;44(10):3883–3888. [11] Tarasov A, Borzenkov I, Milekhina E, Mysyakina I, Belyaev S. Utilization of H2 O2 as the oxygen source by bacteria of the genera Pseudomonas and Rhodococcus. Microbiology. 2004;73(4):392–397.
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[12] Mukherjee AK, Bordoloi NK. Biodegradation of benzene, toluene, and xylene (BTX) in liquid culture and in soil by Bacillus subtilis and Pseudomonas aeruginosa strains and a formulated bacterial consortium. Environ Sci Pollut R. 2012;19(8):3380–3388. [13] Shim H, Hwang B, Lee SS, Kong SH. Kinetics of BTEX biodegradation by a coculture of Pseudomonas putida and Pseudomonas fluorescens under hypoxic conditions. Biodegradation. 2005;16(4):319–327. [14] Horst F, Rueda E, Ferreira M. Activity of magnetiteimmobilized catalase in hydrogen peroxide decomposition. Enzyme Microb Tech. 2006;38(7):1005–1012. [15] Amaretti A, di Nunzio M, Pompei A, Raimondi S, Rossi M, Bordoni A. Antioxidant properties of potentially probiotic bacteria: in vitro and in vivo activities. Appl Microbiol Biotechnol. 2013;97(2):809–817. [16] Zhu BF, Xu Y, Fan WL. High-yield fermentative preparation of tetramethylpyrazine by Bacillus sp. using an endogenous precursor approach. J Ind Microbiol Biot. 2010;37(2):179–186. [17] Du Y, Zhou M, Lei L. The role of oxygen in the degradation of p-chlorophenol by Fenton system. J Hazard Mater. 2007;139(1):108–115. [18] Aminian M, Nabatchian F, Vaisi-Raygani A, Torabi M. Mechanism of Coomassie brilliant blue G-250 binding to cetyltrimethylammonium bromide: an interference with Bradford assay. Anal Biochem. 2012;434(2):287–291. [19] Shukla AK, Vishwakarma P, Upadhyay SN, Tripathi AK, Prasana HC, Dubey SK. Biodegradation of trichloroethylene (TCE) by methanotrophic community. Bioresour Technol. 2009;100(9):2469–2474. [20] Tartakovsky B, Manuel MF, Guiot S. Trichloroethylene degradation in a coupled anaerobic/aerobic reactor oxy-
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genated using hydrogen peroxide. Environ Sci Technol. 2003;37(24):5823–5828. Azcón R, del Carmen Perálvarez M, Roldán A, Barea JM. Arbuscular mycorrhizal fungi, Bacillus cereus, and Candida parapsilosis from a multicontaminated soil alleviate metal toxicity in plants. Microb Ecol. 2010;59(4):668–677. Zhang SG, Han SY, Yang WH, Wei HL, Zhang M, Qi LW. Changes in H2 O2 content and antioxidant enzyme gene expression during the somatic embryogenesis of Larix leptolepis. Plant Cell Tiss Org. 2010;100(1):21–29. Liang HX, Chen AH, Ding C, Li ZX. The response of antioxidant enzyme and ATPase in bacteria exposed to 1, 2dichlorobenzene. Adv Mater Res. 2013;807–809:680–683. Futamata H, Nagano Y, Watanabe K, Hiraishi A. Unique kinetic properties of phenol-degrading Variovorax strains responsible for efficient trichloroethylene degradation in a chemostat enrichment culture. Appl Environ Microbiol. 2005;71(2):904–911. Omsland A, Miranda KM, Friedman RL, Boitano S. Bordetella bronchiseptica responses to physiological reactive nitrogen and oxygen stresses. FEMS microbiol lett. 2008;284(1):92–101. Lu Z, Sang L, Li Z, Min H. Catalase and superoxide dismutase activities in a Stenotrophomonas maltophilia WZ2 resistant to herbicide pollution. Ecotox Environ Safe. 2009;72(1):136–143. Yi Z, Hwa TJ. Co-metabolic degradation activities of trichloroethylene by phenol-grown aerobic granules. J Biotechnol. 2012;162(2–3):274–282. Futamata H, Harayama S, Watanabe K. Group-specific monitoring of phenol hydroxylase genes for a functional assessment of phenol-stimulated trichloroethylene bioremediation. Appl Environ Microbiol. 2001;67(10):4671–4677.
Environmental Technology Publication details, including instructions for authors and subscription information: http://www.tandfonline.com/loi/tent20
Aerobic biodegradation of trichloroethylene and phenol co-contaminants in groundwater by a bacterial community using hydrogen peroxide as the sole oxygen source a
a
ab
c
d
a
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Hui Li , Shi-yang Zhang , Xiao-li Wang , Jie Yang , Ji-dong Gu , Rui-li Zhu , Ping Wang , a
Kuang-fei Lin & Yong-di Liu
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State Environmental Protection Key Laboratory of Environmental Risk Assessment and Control on Chemical Process, State Key Laboratory of Bioreactor Engineering, School of Resources and Environmental Engineering, East China University of Science and Technology, Shanghai 200237, People's Republic of China b
School of Biological Engineering, East China University of Science and Technology, Shanghai 200237, People's Republic of China c
Research Institute of Wastes and Soil Remediation, Shanghai Academy of Environmental Sciences, Shanghai 200233, People's Republic of China d
School of Biological Sciences, Swire Institute of Marine Science, The University of Hong Kong, Pokfulam Road, Hong Kong SAR, People's Republic of China Accepted author version posted online: 15 Sep 2014.Published online: 22 Sep 2014.
To cite this article: Hui Li, Shi-yang Zhang, Xiao-li Wang, Jie Yang, Ji-dong Gu, Rui-li Zhu, Ping Wang, Kuang-fei Lin & Yong-di Liu (2014): Aerobic biodegradation of trichloroethylene and phenol co-contaminants in groundwater by a bacterial community using hydrogen peroxide as the sole oxygen source, Environmental Technology, DOI: 10.1080/09593330.2014.957730 To link to this article: http://dx.doi.org/10.1080/09593330.2014.957730
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Environmental Technology, 2014 http://dx.doi.org/10.1080/09593330.2014.957730
Aerobic biodegradation of trichloroethylene and phenol co-contaminants in groundwater by a bacterial community using hydrogen peroxide as the sole oxygen source Hui Lia,† , Shi-yang Zhanga,† , Xiao-li Wanga,b , Jie Yangc , Ji-dong Gud , Rui-li Zhua , Ping Wanga,b , Kuang-fei Lina and Yong-di Liua∗
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a State Environmental Protection Key Laboratory of Environmental Risk Assessment and Control on Chemical Process, State Key Laboratory of Bioreactor Engineering, School of Resources and Environmental Engineering, East China University of Science and Technology, Shanghai 200237, People’s Republic of China; b School of Biological Engineering, East China University of Science and Technology, Shanghai 200237, People’s Republic of China; c Research Institute of Wastes and Soil Remediation, Shanghai Academy of Environmental Sciences, Shanghai 200233, People’s Republic of China; d School of Biological Sciences, Swire Institute of Marine Science, The University of Hong Kong, Pokfulam Road, Hong Kong SAR, People’s Republic of China
(Received 26 February 2014; final version received 1 August 2014 ) Trichloroethylene (TCE) and phenol were often found together as co-contaminants in the groundwater of industrial contaminated sites. An effective method to remove TCE was aerobic biodegradation by co-metabolism using phenol as growth substrates. However, the aerobic biodegradation process was easily limited by low concentration of dissolved oxygen (DO) in groundwater, and DO was improved by air blast technique with difficulty. This study enriched a bacterial community using hydrogen peroxide (H2 O2 ) as the sole oxygen source to aerobically degrade TCE by co-metabolism with phenol in groundwater. The enriched cultures were acclimatized to 2–8 mM H2 O2 which induced catalase, superoxide dismutase and peroxidase to decompose H2 O2 to release O2 and reduce the toxicity. The bacterial community could degrade 120 mg/L TCE within 12 days by using 8 mM H2 O2 as the optimum concentration, and the TCE degradation efficiency reached up to 80.6%. 16S rRNA gene cloning and sequencing showed that Bordetella, Stenotrophomonas sp., Sinorhizobium sp., Variovorax sp. and Sphingobium sp. were the dominant species in the enrichments, which were clustered in three phyla: Alphaproteobacteria, Betaproteobacteria and Gammaproteobacteria. Polymerase chain reaction detection proved that phenol hydroxylase (Lph) gene was involved in the co-metabolic degradation of phenol and TCE, which indicated that hydroxylase might catalyse the epoxidation of TCE to form the unstable molecule TCE-epoxide. The findings are significant for understanding the mechanism of biodegradation of TCE and phenol co-contamination and helpful for the potential applications of an aerobic bioremediation in situ the contaminated sites. Keywords: chlorinated hydrocarbon; phenol; aerobic biodegradation; hydrogen peroxide; bacterial community
1. Introduction Trichloroethylene (TCE) is a common chlorinated solvent used in industrial processes, and is a widespread contaminant present in soil and groundwater because of improper disposal practices for spent solvents in the past.[1] Meanwhile, amounts of phenols and their close relatives were also disposed from chemical industries, which resulted in the co-contaminations of TCE and phenols distributed in many chemical industrial polluted sites. Biological treatments are considered to be the most effective for the remediation of TCE and phenol co-contaminants because they allow efficient and complete removal from the environment. The TCE biodegradation pathway involves either reductive dechlorination under anaerobic conditions or co-metabolism under aerobic conditions. However, more toxic chemicals such as dichloroethylene and vinyl chloride are generated during the anaerobic process.[2] On
*Corresponding author. Email: [email protected] † Shi-yang Zhang and Hui Li contributed equally to this article. © 2014 Taylor & Francis
the other hand, TCE is partially mineralized to non-toxic materials such as water, carbon dioxide and chlorine ions under aerobic conditions.[3] Although bacterial growth cannot be sustained using TCE as the carbon and energy source, some bacteria utilizing various growth substrates, including methane, phenol, toluene and ammonia, are capable of utilizing TCE as the secondary substrate via cometabolism.[4–6] Groundwater contaminated with TCE has been treated in situ via co-metabolic biodegradation by injecting phenol or toluene.[7] Furthermore, the TCE co-metabolic pathway involving the Lph gene is well established. Lph catalyses the conversion of phenol to catechol and simultaneously co-metabolizes TCE. Chen et al. reported the aerobic degradation process of TCE by phenol-oxidizing bacteria, Pseudomonas putida, which showed a degradation rate of 0.015 mg TCE/mg dry weight/h.[8] Moreover, TCE bioremediation studies
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at contaminated sites and in groundwater have been conducted by adding advantageous bacteria, including the common phenol-utilizing bacteria.[9] Because microorganisms need dissolved oxygen (DO) as the electron acceptor during aerobic biodegradation, the absence of DO in the groundwater renders the aerobic degradation of pollutants difficult. The available oxygen in the groundwater environment could be increased by introducing oxygen into the aquifer by sparging with air, pure oxygen or ozone and by the addition of H2 O2 or oxygen-releasing compounds.[10,11] Use of H2 O2 is a cost-effective approach that is attracting increased attention, because it is soluble in water and provides unlimited amounts of oxygen for biodegradation. Mukherjee et al. have shown that the accretion of 0.01% (v/v) H2 O2 as an additional source of oxygen in the culture medium containing Bacillus subtilis DM-04 and Pseudomonas aeruginosa strains remarkably enhanced the biodegradation rate of benzene, toluene and three xylene isomers under oxygen-limited conditions.[12] However, H2 O2 showed an inhibitory or toxic effect on cell growth with a longer lag phase and slightly lower specific growth rate. Numerous studies have revealed that H2 O2 is toxic to some microorganisms at concentrations as low as 0.003%.[13] Despite the apparent contradiction of using a potential biological toxicant to facilitate aerobic biodegradation, subsurface bioremediation by supplying H2 O2 has been continued on the premise that H2 O2 toxicity would not damage the biodegradation process. Fortunately, aerobic microorganisms have adaptive mechanisms to overcome the adverse effects of external exposure to H2 O2 . Aerobic microorganisms possess catalase (CAT), which decomposes H2 O2 ; superoxide dismutase (SOD), which can remove superoxide ions (O2− ) generated enzymatically; and peroxidase (POD), which uses H2 O2 to oxidize organic compounds.[14] CAT and SOD function cooperatively for protecting microorganisms against H2 O2 toxicity, whereas POD plays a secondary role in the defence mechanism. Increased antioxidase-specific activity was considered to be an indicator of adaptation to H2 O2 under elevated H2 O2 conditions. H2 O2 induces CAT in nearly all aerobic organisms such as Bacillus sp. and Serratia marcescens and induces SOD in Lactococcus spp. and Streptococcus thermophilus.[15,16] H2 O2 has been shown to markedly affect the survival of microorganisms; however, relatively little is known regarding the structure of a microbial community grown in the presence of H2 O2 . This study aimed to obtain a bacterial community that degraded TCE and phenol co-contaminants in the presence of phenol and H2 O2 . An aerobic enrichment was isolated from a site that was exposed to long-term chlorinated hydrocarbon and phenol pollution, and then the enrichment was acclimated to high concentrations of H2 O2 that served as an oxygen source. In addition, the antioxidase activities of the bacteria were determined to confirm
their adaptability to H2 O2 . Finally, the dominant species of the bacterial community were determined, and the mechanism of TCE and phenol co-metabolism was clarified when H2 O2 was used as sole oxygen source.
2. Materials and methods 2.1. Enrichment of the microbial community A soil sample was obtained from a TCE and phenol co-contaminated site with a depth of 0–1.5 m (Pudong, Shanghai, China). For the aerobic culture, 5 g of the soil sample was added to a 125-mL serum bottle containing 50 mL of chloridion-free MCl medium (Na2 HPO4 7.0 g, KH2 PO4 3.0 g, NH4 H2 PO4 2.52 g, Ca(NO3 )2 ·4H2 O 0.024 g, MgSO4 ·7H2 O 0.246 g, H2 O 1000 mL, pH 7.0), followed by the addition of 5 mg/L TCE and 80 mg/L phenol as co-metabolic substances. The serum vials were sealed with butyl stoppers, which were crimped with aluminium caps. Autoclaved vials prepared as above were used as controls. The vials were then cultured at 30°C and shaken at 120 rpm. The enriched cultures were transferred 4–5 times and, simultaneously, the concentration of TCE was constantly increased. Subsequently, 20 mL of the previously mentioned well-grown mixed culture was inoculated into the vials as described above, but amended with 80 mL MCl medium. The resultant solution was cooled under a stream of nitrogen in order to remove oxygen, and then 20 mg/L TCE, 80 mg/L phenol and 2 mM H2 O2 were added. The enriched cultures were acclimatized to H2 O2 by gradually increasing the concentrations of H2 O2 (from 2 to 8 mM in 2 months). The effect on TCE biodegradation (concentration, 20–120 mg/L) in the presence of H2 O2 was also assessed in this study. Cell growth was determined against the control by measuring the optical density at 600 nm (OD600 ).
2.2. TCE degradation analysis TCE biodegradation was monitored using gas chromatography (GC). For this, a 1 mL sample was obtained from each vial by using a sterile syringe and mixed with 1 mL n-hexane. After the sample was oscillated and extracted, 1 μL organic phase was analysed using GC by using Agilent A7890 equipped with an electron capture detector. The injector, oven and detector temperatures were set at 240°C, 80°C and 260°C, respectively. Nitrogen (flow rate, 20 mL/min) was used as the carrier gas. Samples were injected at a split rate of 5:1, and the retention time of TCE was approximately 5.1 min. The concentration of TCE degradation was corrected for the volatilized amount of TCE by using the following equation: TCEc = TCEM + rvol.TCE × t, where TCEC , TCEM , rvol.TCE and t are corrected TCE concentration (ppm), measured TCE concentration (ppm), TCE volatilization rate (ppm/day) and time in days, respectively.
Environmental Technology 2.3. H2 O2 and DO analyses The titanium oxalate method was used to spectrophotometrically quantify the H2 O2 levels.[17] The absorbance at 400 nm was measured using a UV spectrophotometer (UV 1800; SHIMADZU). The assay was used for H2 O2 concentration ranging from 0 to 12 mM. DO of the samples was detected using a Clark-type electrode (YSI 550-12 Oxygen Meter; Yellow Springs). A three ml sample was injected to the 10 mL-reaction vessel and the probe was inserted into the liquid. The reaction vessel was sealed with Teflon film. A wait of 10 min ensued thereafter by setting the electronic timer, to get a stable reading and write down the measurement of conductivity.
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2.4.
Key enzyme activity assays
The protein concentration of the bacterial suspension was determined using the Coomassie blue method.[18] Coomassie Brilliant Blue G-250 and bovine serum albumin were used for the generation of standard curves. The related enzyme activities were measured by preparing crude extracts. For this, 10 mL culture was harvested by centrifugation (4000 g, 10 min, 4°C), and the supernatant was removed. Subsequently, the cells and 10 mL phosphate buffer solution were mixed using a vortex mixer for a few seconds. The washing process was repeated three times. Next, the cell suspensions were sonicated for 10 min at 45-s bursts at an output level of 6 by using a cup horn attachment, and then the centrifuge tubes were maintained below 4°C. Cell debris was removed by centrifugation at 12,000 g; the supernatant was used for enzyme assays. CAT activity was assayed as the rate of decrease of the absorbance resulting from the disappearance of 0.01 mol/L H2 O2 solution (0.05 mol/L phosphate buffer; pH 7.0; 25°C) at 240 nm. One unit of enzyme activity is the amount of CAT that degrades 1 mmol of H2 O2 per minute. ® The SOD activity was assayed using an SOD assay-WST kit (Dojindo, Japan). The enzyme and WST-1 working solutions and crude extracts were prepared according to the manufacturer’s instruction. After all the solutions were added to 96-well microplates as described by the manufacturer, the microplates were incubated at 37°C for 20 min. Subsequently, the absorbance was measured at 450 nm by using a microplate reader (Sunrise, Tecan, Austria). One unit of SOD activity was defined as the amount of enzyme in a sample solution causing 50% inhibition (IC50 ) of the rate of tetrazolium reduction. POD activity was assayed using the method in which the rate of formation of coloured product resulting from the oxidation of guaiacol is measured at 460 nm. The buffer contained 0.2 mol/L phosphate buffer (pH 7.8). The reaction mixture contained 0.1 mol/L phosphate buffer (pH 6.0), 4 mmol/L H2 O2 and 0.6 mg/mL guaiacol. For the assay, 1 mL of the enzyme extract was mixed with 3 mL of the reaction mixture. One unit of enzyme activity was defined as change rate in the absorbance of 45 s.
2.5.
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DNA extraction, cloning and sequencing
After domestication, 10 mL of aerobic medium exposed to H2 O2 was filtered through a sterile 0.22-μm membrane. DNA from the cells was extracted using a bead beating technique (Biospec, Shanghai, China) by using a Fast DNA spin kit for DNA extraction (Abigen, Beijing, China). Polymerase chain reaction (PCR) amplifications were performed using a 25-μL reaction mixture that contained 0.5 μL of template, 0.5 μL of each primer (10 mmol/L; 27F: 5 -AGAGTTTGATCCTGGCTCAG-3 ; 1492R: 5 TACGGTTACCTTGTTACGACTT-3 ), 12.5 μL Taq mix and 11 μL of ddH2 O. The PCR was carried out using the following cycling conditions: 1 cycle of initial denaturation at 94°C for 5 min, denaturation at 94°C for 1 min, annealing at 52°C for 30 s, extension at 72°C for 1 min; 30 cycles of denaturation at 94°C for 1 min, annealing at 52°C for 30 s, extension at 72°C for 1 min; 1 cycle of re-extension at 72°C for 10 min and a final storage at 4°C. The PCR was performed using a Mastercyle gradient (Eppendorf, US). The PCR products were purified using the DNA purification kit (V-gene; Shanghai, China) and then ligated into the pMD 19-T vector system according to the manufacturer’s instructions (TaKaRa, Dalian, China). The ligated plasmids were transformed into Escherichia coli DH5α. The insertion of 16S rRNA gene was screened and retrieved by PCR amplification by using the primer set of M13-47 (5 -CGCCAGGGTTTTCCCAGTCACGAC-3 ) and RVM (5 -GAGCGGATAACAATTTCACACAGG-3 ). The positive clones were sequenced (BGI, Beijing, China). After the sequences were edited and checked manually, the identity of the typical 16S rRNA gene sequences was checked against published sequences in the GenBank database of National Center for Biotechnology Information (http://www.ncbi.nlm.nih.gov) by using the BLAST software. Phylogenetic trees were constructed using Clustal X (1.8) and Molecular Evolutionary Genetics Analysis (MEGA, version 5.05). Robustness of individual branches was estimated by bootstrapping based on 1000 replications.
2.6.
Analysis of the phenol hydroxylase gene involved in TCE co-metabolism
The isolated DNA was used for the amplification of the phenol hydroxylase (Lph) gene that is known to be involved in the phenol-TCE co-metabolic pathway. The gene was amplified using the following primers sets: Lphf (5 -CGCCAGAACCATTTATCGATC-3 ) and Lphr (5 -AGGCATCAAGATCACCGACTG-3 ).
2.7. Nucleotide sequence accession numbers The GenBank accession numbers for bacterial 16S rRNA gene sequences are KF926834 and KF926836–KF926839 and that for the functional gene is KJ004664.
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H. Li et al. the degradation efficiency increased up to 80.6% with the degradation rate of 8.06 ppm/day when the medium were supplied with 8 mmol/L H2 O2 . The bacterial community showed increased tolerance to TCE from 20 to 120 mg/L and the degradation capacity of TCE also increased, which might have contributed to treat the in situ TCEcontaminated groundwater. Meanwhile, the enrichment was able to degrade phenol at H2 O2 concentrations from 2 to 8 mM (Figure 2). Phenol was degraded after 2–4 days and most of phenol biodegradation was achieved when H2 O2 was added to the medium. Finally, the enrichment almost completely degraded 220 mg/L phenol when 8 mM H2 O2 served as the oxygen source. Previous study documented that only the lab-scale biodegradation of TCE occurred under sufficient oxygen environment. Shukla et al. [19] isolated a methanotrophic community from an uncontaminated tropical soil to degrade TCE. The community degraded 50% TCE in 8 days from an initial concentration of 100 ppm.[19] Tartakovsky et al. [20] showed that the performance of a TCE-degrading microbial consortium was improved by using H2 O2 as an oxygen source in a biofilm reactor. The consortium degraded 18 mg/L TCE per day with supplementation of 0.7 g H2 O2 per day).[20] Compared to the previous study, our findings showed a higher TCE biodegradability as well as H2 O2 tolerance in the phenolloaded media. Few studies have reported the degradation of higher concentrations of TCE by phenol-utilizing bacteria in the presence of H2 O2 .
3.2. Figure 1. Variation of TCE concentration and OD600 during the bacterial community acclimating 2–8 mM H2 O2 . The numbers on the top represent the concentration of H2 O2 in each period.
Decomposition of H2 O2 to produce DO Figure 3 showed the variation of H2 O2 concentration and DO during the bacterial community acclimation. When 2 mM H2 O2 was introduced into the medium, the concentration of DO reached 6 mg/L, which indicated that
8 mM H2O2
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Concentration of Phenol
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CPhenol /mg/L
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3. Results and discussion 3.1. Acclimating bacterial community to degrade TCE and phenol co-contaminants The enrichment was acclimated to degrade the cocontaminants of 120 mg/L TCE and 220 mg/L phenol in the presence of increasing H2 O2 concentrations at 2–8 mM in 2 months (Figures 1 and 2). At low concentrations of H2 O2 (2 mmol/L), the cell grew rapidly after a short lag phase. Cells continued to grow and reached the highest OD600 when 8 mmol/L H2 O2 was added. Nevertheless, OD600 reached 0.1 after the addition of 12 mmol/L H2 O2 (data not shown). The enriched medium possessed the ability to biodegrade 20–120 mg/L TCE after the addition of H2 O2 . Initially, 20 mg/L TCE was almost reduced by 68% within 6 days when 2 mmol/L H2 O2 was added. Finally,
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Figure 2. Variation of phenol concentration during the bacterial community acclimating 2–8 mM H2 O2 . The numbers on the top represent the concentration of H2 O2 in each period.
Figure 3. Variation of H2 O2 concentration and DO during the bacterial community acclimation. The numbers on the top represent the concentration of H2 O2 in each period.
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Environmental Technology H2 O2 was decomposed to release the DO. On the eighth day, the concentration of DO was considerably lower than before, indicating that the bacteria could adapt to 2 mmol/L H2 O2 and extract higher oxygen from H2 O2 . Similarly, the observed DO values after treatment with 8 mmol/L H2 O2 increased to 20 mg/L, and then slowly dropped to 5 mg/L finally. H2 O2 was rarely detected in the entire culture process, but H2 O2 decomposition became considerably slow and gas production increased remarkably after the addition of 12 mM H2 O2 (data not shown). The results indicated that DO would not be the limiting factor for TCE co-metabolism. Supplementation of H2 O2 in increasing doses to the microorganisms allowed them to adapt to H2 O2 . The DO concentration decreased in every phase, indicating that microorganisms need to adapt to H2 O2 . The optimized concentration of H2 O2 was 8 mM in view of the tolerance of H2 O2 and the available oxygen production. 3.3.
Tolerance mechanism of the bacterial community to H2 O2
The activities of the three antioxidases, CAT, SOD and POD, of the enriched culture were influenced by various concentrations of H2 O2 (Table 1). The CAT levels in the presence of 2–8 mmol/L H2 O2 continued to increase and were greater than the value after treatment with 12 mmol/L H2 O2 . Similar to CAT, SOD activity was the highest at 8 mmol/L H2 O2 . The SOD activity at 12 mmol/L H2 O2 was slightly lower than that detected after treatment with other concentrations of H2 O2 , similar to that noted with CAT. The POD activity varied between 0.045 and 0.138 U/mg for the four concentrations of H2 O2 . The difference might seem minor, but the highest POD activity was also observed when 8 mmol/L H2 O2 was added. These three enzymes were often used to evaluate the antioxidant property of bacteria and plants.[21,22] Our data suggest that CAT, SOD and POD activities increased after treatment with 2–8 mmol/L H2 O2 , indicating the tolerance and adaption of the bacteria in the enriched medium to H2 O2 . Nevertheless, the activities of these three enzymes were inhibited remarkably after treatment with 12 mmol/L H2 O2 , which indicated that the enriched culture could not withstand the supplementation of 12 mmol/L H2 O2 . In the previous studies, CAT and SOD were also considered as the main antioxidant enzymes in aerobic organisms.[23] Table 1.
The increases in antioxidase activities of the enriched culture after treatment with 2–8 mmol/L H2 O2 might be attributed to a shift in the bacterial community towards members having intrinsically higher antioxidant activity levels, or to the induction of antioxidases by specific members of the microbial population. The bacterial consortium in this study overcame the toxic effects of H2 O2 by inducing antioxidases when grown in the H2 O2 -supplemented environment. 3.4. The bacterial community structure The community structure of the acclimated enrichment was investigated by developing a clone library, and a total of 30 positive clones from bacterial 16S rRNA gene clone libraries were clustered into five operative taxonomic units (OTUs) with 97% similarity by using Muthor 1.32. The rarefaction curve became flat to the right, indicating that a reasonable number of individuals of the community were sampled. The phylogenetic tree elucidated the dominant species in the enrichment distributed in the three major groups: Betaproteobacteria (19/30, 63.33%, 2 OTUs), Alphaproteobacteria (6/30, 20%, 2 OTUs) and Gammaproteobacteria (5/30, 16.67%, 1 OTU; Figure 4). Clh4 showed high similarity (99%) to Bordetella sp. KP22 (AB015607), which was isolated from the polluted soil with TCE contamination and showed strong TCE-degrading ability. Clh2 exhibited 98% similarity to aerobic para-nitrophenol degrader Stenotrophomonas sp. CERAR5 (JF346409), which was found in a TCE-contaminated site. Clh3 exhibited 96% similarity to Sinorhizobium sp. c37 (AB167207), which was found to be responsible for TCE degradation in a phenol-oxidizing culture.[24] Clh1 exhibited 99% similarity to a bacterium clone ccslm224 (AY133088), which was isolated from a TCE-polluted site. Clh5 exhibited 99% similarity to Sphingobium sp. XTT-3 (JN120899) that could degrade phenol. Bordetella sp. Clh4 was the most dominant species in the enrichment, and Bordetella are known to produce SOD and CAT to detoxify superoxide and H2 O2 .[25] Further, some Stenotrophomonas sp. were pollutant-degrading bacteria having high antioxidant activity.[26] Analysis of the dominant species explained why the enrichment showed considerably higher TCEdegrading activities and was adapted to H2 O2 during the acclimation.
The antioxidase activities after treatment with different concentrations of H2 O2 .
CAT per U/mg protein SOD per U/mg protein POD per U/mg protein
2 mmol/L
4 mmol/L
8 mmol/L
12 mmol/L
2.463 ± 0.317 0.492 ± 0.108 0.047 ± 0.011
2.534 ± 0.248 0.522 ± 0.131 0.065 ± 0.014
3.725 ± 0.286 0.916 ± 0.129 0.138 ± 0.012
0.932 ± 0.195 0.239 ± 0.093 0.045 ± 0.012
Results represent mean ± SD for three independent experiments. p < .05.
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H. Li et al. 100
Clh4 Bordetella sp. KP22 (AB015607)
68
Oxalobacteraceae bacterium HTCC315 (AY429715) Burkholderia cepacia strain AW201 (EF373556)
99
Clh1
92
clone ccslm224 (AY133088) Variovorax paradoxus MBIC3839 (AB008000)
100 80
Clh2 Stenotrophomonas sp. CERAR5 (JF346409) Stenotrophomonas maltophilia strain:c309 (AB167179) Stenotrophomonas maltophilia strain PM102 (JQ797560) 100
Clh5 Sphingobium sp. XTT-3 (JN120899)
66 91
clone 290 (DQ158105) Sphingomonas sp. FG03 (EU784670) Clh3
100
Rhizobium sp. c21 (AB167200)
99
Alphaproteobacteria
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97
Gammaproteobacteria
clone GPT1a (AY706440)
96
100
Betaproteobacteria
71
Sinorhizobium sp. c37 (AB167207)
81 99
Sinorhizobium sp. c93 (AB167234)
0.02 Figure 4. Phylogenetic tree of the V6 partial sequence of bacterial 16S rDNA from the bacteria grown in a culture medium supplemented with 8 mmol/L H2 O2 is shown in boldface, and the referred sequences in the EMBL database with the putative divisions are listed to the right. The topology was calculated using the neighbour-joining method. Bootstrap values n = 1000 replicates. More than 50% of the values were reported near the corresponding nodes. The scale bar represents 0.02 nuclear acid substitutions per nucleotide position.
3.5.
Mechanism of TCE and phenol co-metabolism by using H2 O2 as sole oxygen source Whether TCE degradation by the bacterial community grown in a H2 O2 -supplemented environment proceeded via the known pathways [27] was determined by performing PCR analysis of key genotypes by using primer sets for gene sequences encoding Lph, which is involved in the epoxidation of TCE. The PCR products showed the expected size for the Lph gene (684 bp) for the test sample; further, there was no expression found in the negative samples (Figure 5). The sequencing and BLAST analysis revealed that the sequence had 85% similarity to the Lph genes involved in the TCE-degradation pathway. The epoxidation of TCE was the key metabolic reaction, and many published strains were known to effectively degrade TCE via this pathway, such as Ralstonia eutropha, Variovorax sp. and Pseudomonas sp.[28] The present study showed that the Lph gene was actively involved in the TCE co-metabolism under an H2 O2 -supplemented condition. The results suggested that the mechanism of degrading
Figure 5. PCR product analogues of the indicated catabolic genotypes obtained using DNA extracts of bacteria grown in a medium containing 8 mmol/L H2 O2 by using primer set Lph.
Environmental Technology
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Figure 6. The mechanism of TCE and phenol co-metabolism in the presence of H2 O2 .
TCE was improved by adding H2 O2 (Figure 6). H2 O2 induced antioxidase enzymes of the bacteria to release O2 ; further, the antioxidases relieved the toxicity to the bacteria caused by H2 O2 and O2− . In addition, the bacterial community was stimulated by phenol to produce Lph enzymes which catalysed the reaction for phenol oxidation. Because the Lph enzyme was an iosenzyme, it catalysed the reactions for O2 and TCE forming TCE-epoxide which was a chemically unstable molecule. The TCE-epoxide spontaneously decomposed into readily degradable compounds, and finally they were degraded to CO2 , H2 O and chloride ions. 4.
Conclusions
In this study, a bacterial community was enriched from TCE-contaminated soil to degrade TCE in the presence of H2 O2 . The H2 O2 -tolerant mechanism of the enrichment involved increasing the activities of antioxidant enzymes. The bacterial community included: Bordetella, Stenotrophomonas, Sinorhizobium, Variovorax and Sphingobium. The dominant species belonged to Bordetella and Stenotrophomonas. The bacterial community induced antioxidase enzymes to release O2 to improve the DO in the system. The Lph enzymes catalysed O2 and TCE to form TCE-epoxide which was readily degradable compounds. This study contributed to the potential applications of an aerobic process for in situ bioremediation of TCE-contaminated groundwater. Acknowledgements This work was supported jointly by National Natural Science Foundation of China (51378208, 41273109, 41003031), Specialized Research Fund for the Doctoral Program of Higher Education (20110074130002), Shanghai Rising-Star Program (12QA1400800), Fok Ying Tung Education Foundation (141077), Program for New Century Excellent Talents in University (NCET-13-0797), Innovation Program of Shanghai Municipal Education Commission (14ZZ059), Fundamental Research Funds for the Central Universities (222201313008). We also
would like to thank the anonymous referees for their helpful comments on this paper.
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