Changes in bacterial community of anthracene bioremediation in municipal solid waste composting soil

Changes in bacterial community of anthracene bioremediation in municipal solid waste composting soil first-author corresp ‡Corresponding Author ‡Corresponding Author †E-mail:xiesg@pku.edu.cn †E-mail:xiesg@pku.edu.cn xiesg@pku.edu.cn College of Environmental Sciences and Engineering, Peking University, Beijing 100871, China College of Environmental Sciences and Engineering, Peking University, Beijing 100871, China Polycyclic aromatic hydrocarbons (PAHs) are common contaminants in a municipal solid waste (MSW) composting site. Knowledge of changes in microbial structure is useful to identify particular PAH degraders. However, the microbial community in the MSW composting soil and its change associated with prolonged exposure to PAHs and subsequent biodegradation remain largely unknown. In this study, anthracene was selected as a model compound. The bacterial community structure was investigated using terminal restriction fragment length polymorphism (TRFLP) and 16S rRNA gene clone library analysis. The two bimolecular tools revealed a large shift of bacterial community structure after anthracene amendment and subsequent biodegradation. Genera Methylophilus , Mesorhizobium , and Terrimonas had potential links to anthracene biodegradation, suggesting a consortium playing an active role. Anthracene Microbial community Biodegradation Bioremediation Introduction 1 With rapid urbanization, municipal solid waste (MSW) constitutes an immediate and serious environmental problem in many developing countries. In many Chinese urban and suburban regions, a great amount of collected MSW is dumped on land in a more or less uncontrolled manner. Leachate from MSW composting site may contain a large number of xenobiotic organic compounds (Baun et al., 2004 2004 ; Xu et al., 2008 2008 ). Therefore, inadequate MSW disposal can lead to the contamination of soil, and surface and ground waters in the vicinity of the site. Polycyclic aromatic hydrocarbons (PAHs) are a group of chemical compounds consisting of carbon and hydrogen, whose ubiquitous existence in the various environments has aroused great concerns. Leachate from an unmanaged MSW composting site can cause heavy PAH contamination in soil in the vicinity of the site (Han et al., 2009 2009 ). Microbial degradation plays the primary role in PAH reduction in contaminated sites (Yuan et al., 2001 2001 ; Yousefi Kebria et al., 2009 2009 ). The knowledge about the microbial community structure in MSW composting soil is very important for development of bioremediation technology. Numerous studies have investigated the bacterial community structure in contaminated aquifers near MSW composting sites (Röling et al., 2000 2000 ; 2001 2001 ; Tian et al., 2005 2005 ; Yu et al., 2010 2010 ). However, these previous works have not investigated the links of PAHs to the bacterial communities. Also, little is known about the microbial community in MSW soil. The concentration of PAHs in soils near MSW compositing site usually increases with time (Han et al., 2009 2009 ). Unfortunately, very few investigations have been undertaken to assess the impact of prolonged PAH contamination or PAH amendment on soil microbial community (Muckian et al., 2009 2009 ). As a consequence, the microbial community in the MSW composting soil and its change associated with prolonged exposure to PAHs and subsequent biodegradation remain largely unknown. Anthracene has been listed as one of the priority environmental pollutants by the United States Environmental Protection Agency. Anthracene is also a model compound for PAH degradation studies (Müncnerová and Augustin, 1994 1994 ; Krivobok et al., 1998 1998 ). An average level of 1 208.5 µg/kg total PAHs (including 88.2 µg/kg anthracene) has even been observed in an MSW composting soil (Han et al., 2009 2009 ). Therefore, the aim of current study was to investigate the bacterial community in the MSW composting soil, and to explore the change associated with anthracene amendment and subsequent biodegradation. The genera with potential PAH degradation capabilities were also investigated. Materials and methods 1 Microcosm experiment set-up 2 Soil was collected from the vicinity of an unmanaged MSW composting site in Changzhou, a southeast city in China. Following sample collection, soil was air-dried, homogenized, and sieved through a 0.18-mm screen, and stored at 4 °C until use. Anthracene (99%, J & K China Chemical) dissolved in acetone was added to each empty microcosm chamber (150 ml serum bottle) with a total mass of 200 µg anthracene per chamber. After the acetone evaporated, 2 g dry soil was added to each microcosm along with 10 ml phosphate buffered mineral media, as previously described (Mu and Scow, 1994 1994 ). The bottles were sealed with rubber stoppers and aluminum seals. The sterile controls were obtained by autoclaving (121 °C, 1 h) repeatedly (three times). Microcosms were maintained at 1 °C in a refrigerator and stirred periodically (manually stirred for 2 min, every 30 min) for 48-h equilibration. Microcosms were incubated at 25 °C (Day 0). Microcosms were sampled on Days 0, 10, and 15. In each sampling time, three replicate microcosms were used for anthracene analysis and the other three for soil DNA extraction. Anthracene extraction and analysis 2 In order to determine anthracene in solid phase, the water/soil mixture in microcosms was transferred to a 100-ml centrifuge tube. After 10-min centrifugation at 5 000 r/min, the supernatant was collected. A total of 1 g dry soils were extracted three times with 10 ml acetone, using a 300-W ultrasonic processor. The mixture was vigorously shaken and then centrifuged at 5 000 r/min for 10 min. Then 0.2 ml of the supernatant was collected into a gas chromatography (GC) vial and diluted by 0.8 ml methanol (Zhang et al., 2011 2011 ). The mixture was filtered with a 0.45-µm British Nylon 66 syringe filter provided by Tianjin Jingteng Corp., China. The anthracene analysis was conducted according to the standard method (Huang Y. et al., 2010 2010 ). Terminal restriction fragment length polymorphism (TRFLP) analysis 2 Soil DNAs from Samples A (from microcosm on Day 0) and B (from microcosm on Day 15) were used for further molecular analysis. DNA was extracted using the UltraClean DNA extraction kit (Mo Bio Laboratories). Bacterial 16S rRNA genes were amplified using the bacterial primers 27F-FAM (5′-GAG TTT GAT CMT GGC TCA G-3′, 5′ end-labeled with carboxyfluorescein) and 1492R (5′-GGT TAC CTT GTT ACG ACT T-3′). The polymerase chain reaction (PCR) program was as previously described (Cupples and Sims, 2007 2007 ; Zhang et al., 2011 2011 ). A total of 300 ng fluorescently labeled PCR products were purified with QIA quick PCR purification kit (Qiagen Inc., Germany) and digested with Hae III with a 6-h incubation period. Hha I and Msp I were also used to confirm the identity of the greatly enriched fragments. The same purified PCR product was used for digestion with Hae III, Hha I, and Msp I. The clone sequence was first identified, the in silico cut site ( Hae III digest) of which matched the length of the greatly enriched fragment ( Hae III digest). If the clone restriction enzyme ( Hha I or Msp I) cut site predicted from sequence also matched the observed length of one abundant fragment ( Hha I or Msp I digest) in Sample B, the taxonomic identity of the greatly enriched fragment ( Hae III digest) could be confirmed (Cupples and Sims, 2007 2007 ; Zhang et al., 2011 2011 ). The fragment pattern was detected using an ABI 3730 DNA analyzer set at the GeneScan mode together with an internal lane standard (GeneScan-500 LIZ Size Standard, Applied Biosystems). Replicate profiles from separate DNA extractions and PCR reactions for each sample were compared to identify the subset of reproducible fragment sizes. The average area of each reproducible peak was calculated. Terminal fragments smaller than 50 bp or under 200 fluorescent units were excluded from the further analysis. Fragments differing by less than 1 bp length were clustered. The standardized binning criteria used to identify the subset of reproducible peaks were as previously described (Dunbar et al., 2001 2001 ; Mills et al., 2003 2003 ). Ribotype richness ( S ) equals to the total number of distinct fragments in a profile. The Shannon diversity index ( H ) and evenness ( E ) were calculated according to the standard methods (Mills et al., 2003 2003 ; Falk et al., 2009 2009 ). Bray-Curtis similarity index was calculated using PRIMER 5.0 software to evaluate the similarity for two microcosm samples (Clarke and Warwick, 2001 2001 ). Similarity percentage (SIMPER) analysis was further used to identify the fragments that were mainly responsible for the dissimilarity between samples (Clarke and Warwick, 2001 2001 ; Rees et al., 2004 2004 ). DNA cloning and sequencing 2 The PCR conditions were the same as the above-mentioned, except that the forward primer was unlabeled. The PCR products were cloned into pMD19-T vector (TaKaRa Co., Japan) following the manufacturer’s instruction. The white colonies were verified by PCR with primers M13 F (5′-TGT AAA ACG ACG GCC AGT-3′) and M13 R (5′-AAC AGC TAT GAC CAT G-3′). Clones containing an insert of the correct size were bidirectionally sequenced. The 69–72 clones recovered from each sample were successfully sequenced in this study. The partial 16S rRNA gene sequences were deposited with GenBank under accession Nos. HQ015161-HQ015301. Sequences for chimerism were checked using the Chimera Check program available at the Michigan State University (MSU) Center for Microbial Ecology. All clones displaying chimeric profiles were discarded from the further analysis. Sequences that were over 97% similar were grouped into an operational taxonomic unit (OTU) by manual comparison. The representative GenBank sequences to the clones of interest were extracted from the National Center for Biotechnology Information (NCBI) database and included in further phylogenetic analyses using MEGA Version 4.0 (Tamura et al., 2007 2007 ). The neighbor-joining trees were constructed using p -distance with 1 000 replicates to produce bootstrap values. Alignment of the sequences was carried out using ClustalW ( http://www.ebi.ac.uk/clustalw/ http://www.ebi.ac.uk/clustalw/ ). The Ribosomal Database Project (MSU Center for Microbial Ecology) analysis tool “classifier” ( http://rdp.cme.msu.edu/classifier/classifier.jsp http://rdp.cme.msu.edu/classifier/classifier.jsp ) was utilized to assign taxonomic identity (Wang et al., 2007 2007 ). Results 1 Biodegradation 2 In this study, the total amount of anthracene in supernatant was below 3% of that in the solid phase over the whole experiment (data not shown). Anthracene concentration in soil declined greatly after a 15-d experiment period. In contrast, only a limited decline in soil was observed in the autoclaved control, possibly due to aging effect. This confirmed a biological removal mechanism in microcosms (Table 1 1 ). 2 Time (d) 2 Remaining ANT (%) Sterile control with ANT Microcosm with ANT 0 97.5 97.2 10 80.9 44.8 15 77.0 11.4 TRFLP analysis 2 The total average numbers of terminal restriction fragments ( Hae III digest) in Samples A and B were 35 and 29, respectively (Fig. 1 1 ). Shannon diversity index and evenness did not differ much between Samples A and B (Table 2 2 ). However, the low Bray-Curtis similarity (average 38%) reflected a big difference of microbial community structure between Samples A and B. S : ribotype richness; S =total number of bands in profile. H : Shannon diversity index; H =−∑( p i )(log 2 p i ), where p i is the individual peak area; H max =log 2 S . E : evenness; E = H / H max Samples S H E A 35 3.191 0.897 B 29 3.145 0.934 The SIMPER analysis revealed that the differences of TRFLP profiles ( Hae III digest) between Samples A and B were driven primarily by variation in 217, 79, 222, and 258 bp (Table 3 3 ). The 79 bp was abundant in Sample A. These abundant fragments (217, 222, and 258 bp) in Sample B were not detected in Sample A (Fig. 1 1 ), which indicated these fragments were largely enriched with anthracene biodegradation. Mean relative abundance of each fragment ( Hae III digest) as a percentage of total fragment abundance Fragment contribution as a dissimilarity percentage between the two groups Lists are truncated to include only those fragments that contribute no less than 5% to the differences between samples 2 Fragment (bp) 2 Mean relative abundance (%) a a 2 Fragment contribution (%) b b Sample A Sample B 217 0 10 6 79 10 0 6 222 0 8 5 258 0 8 5 Phylogeny 2 In this study, many known phyla (Bacteroidetes, Proteobacteria, Verrucomicrobia, Acidobacteria, TM7, Cyanobacteria, Gemmatimonadetes) were detected, although only Bacteroidetes and Proteobacteria were shared between both samples (Fig. 2 2 ). For Sample A, the 72 recovered clones were distributed across phyla as follows: β-proteobacteria 43.1%, γ-proteobacteria 15.3%, α-proteobacteria 12.5%, unclassified bacteria 8.3%, δ-proteobacteria 6.9%, Gemmatimonadetes 6.9%, Bacteroidetes 4.2%, Acidobacteria 2.8%. However, the major phylum types (10% or more) of 69 clones recovered from the Sample B were γ-proteobacteria (44.9%), β-proteobacteria (23.2%), and Bacteroidetes (14.5%). Compared with Sample A, Bacteroidetes and γ-proteobacteria achieved a large increase in relative abundance (no less than 10%). Unclassified Proteobacteria, Verrucomicrobia, TM7, and Cyanobacteria were also detected in Sample B. These results show a shift of bacterial community structure in MSW composting soil. Clones not classified to any known phylum are included as unclassified bacteria Moreover, the 72 clones in the bacterial library constructed with Sample A could be further divided into 46 OTUs (14 OTUs had two or more clones). However, the 69 clones in library constructed with Sample B were grouped into 39 OTUs (11 OTUs had two or more clones). Therefore, in agreement with the TRFLP analysis, the bacterial community structure reflected by the clone library analysis also indicated a big change with PAH biodegradation. Identity of enriched TRFLP fragments 2 The taxonomic identity of 217, 222, or 258 bp ( Hae III digest) was also investigated. Comparisons of TRFLP fragments and in silico cut sites are presented in Table 4 4 . The taxonomic identity of each enriched fragment is presented in Table 5 5 . The comparisons of TRFLP fragments and in silico cut sites to determine the enriched fragments have also been noted elsewhere (Cupples and Sims, 2007 2007 ; Luo et al., 2009 2009 ). 217, 222, and 258 bp ( Hae III digest) obtained from Sample B were identified as Methylophilus , Mesorhizobium , and Terrimonas , respectively. The numbers of clones classified within genera Methylophilus , Mesorhizobium , and Terrimonas in Sample B were 6, 2, and 7, respectively. However, only one Methylophilus clone and one Terrimonas clone were detected in Sample A. Therefore, similar to TRFLP results (Fig. 1 1 ; Table 3 3 ), the enrichment of the three genera was also demonstrated. Enriched fragment Restriction enzyme TRFLP Sequence data 3 1 Hae III 217 219 Msp I 487 490 Hha I 367 367 4 3 2 Hae III 222 225 Msp I 128 128 Hha I 58 61 4 3 3 Hae III 258 260 Msp I 90 93 Hha I 69 67 Fragment (bp) Phylum Class Order Family Genus 217 Proteobacteria β-proteobacteria Methylophilales Methylophilaceae Methylophilus 222 Proteobacteria α-proteobacteria Rhizobiales Phyllobacteriaceae Mesorhizobium 258 Bacteroidetes Sphingobacteria Sphingobacteriales Chitinophagaceae Terrimonas Phylogenetic tree of two representative sequences in each genus ( Methylophilus , Mesorhizobium , or Terrimonas ) was constructed using MEGA Version 4.0 (Fig. 3 3 ). Clones L5 and L6 were identified as the members of genus Methylophilus . Their three closest cultivated matches were obtained from human organs (AB193725.1, 99% identity), microcystis colonies from Taihu Lake (EU770265.1, 99% identity), and field grown rice cultivars (EU194887.1, 99% identity). Clones L16 and L46 were classified within genus Mesorhizobium . Their three closest cultivated matches were obtained from grassland soil (AB531424.1, 99% identity), legume genus Biserrula growing in Eritrea (FJ827044.1, 99% identity), and root nodules (FJ814715.1, 99% identity). Clones L2 and L4 were identified as Terrimonas species. The two clones illustrated 95% similarity to a Terrimonas lutea sp. species (AB192292.1) isolated from soil. They were also related to several cultured sequences, such as fresh water Flavobacteriaceae species (AB264128.1, 96%), soil Sphingobacteriales species (EU854577.1, 96%), and soil rhizosphere species (AJ252591.1, 95%). Data in parentheses are GenBank accession numbers. Numbers at the nodes indicate the levels of bootstrap support based on neighbor-joining analysis of 1 000 resampled datasets Discussion 1 Considerable attention has been paid to the diversity of indigenous microorganisms capable of degrading pollutants in various environments (Abed et al., 2002 2002 ). Some other profiling techniques, such as community-level physiological profiling, restriction fragment length polymorphism, denaturing gradient gel electrophoresis, and phospholipid fatty acid profiling, have been applied to investigate the composition of bacterial community in contaminated aquifers near MSW composting sites (Röling et al., 2000 2000 ; 2001 2001 ; Tian et al., 2005 2005 ; Yu et al., 2010 2010 ). Among available DNA profiling techniques, TRFLP has been used to monitor structural changes of various soil microbial communities (Dunbar et al., 2000 2000 ; Singh et al., 2006 2006 ; Grant et al., 2007 2007 ). However, to the best of our knowledge, this is the first report to investigate the change of the bacterial community in the MSW composting soil in response to anthracene amendment and subsequent biodegradation. The bacterial community structure usually changes during the bioremediation of PAH-contaminated soils (MacNaughton et al., 1999 1999 ; Vinas et al., 2005 2005 ; Gandolfi et al., 2010 2010 ). Muckian et al. ( 2009 2009 ) have revealed the significant change of the microbial community associated with fluoranthene or phenanthrene amendment and subsequent biodegradation. Piskonen et al. ( 2005 2005 ) have also shown selection in the microbial community as a result of naphthalene biodegradation. In this study, both TRFLP and 16S rRNA gene clone library analysis revealed a big shift of microbial community structure with anthracene biodegradation. The phylogenetic description of the bacterial community structure is also important. Tian et al. ( 2005 2005 ) has reported the dominance of Firmicutes and Fusobacteria in the leachate-contaminated aquifers. Although the literature on the microbial community in MSW composting soil is still lacking, the dominance of Proteobacteria has been observed in the soil/water system for PAH biodegradation (Chang et al., 2007 2007 ), and in aged PAH-contaminated soil (Cébron et al., 2009 2009 ). Our preliminary research indicated that the soil used to construct microcosms contained high concentration of anthracene (about 80 µg/kg). Therefore, the abundance of Proteobacteria in Sample A might be partly attributed to the high level of PAHs in MSW composting soil. Many different bacterial genera previously isolated from PAH-contaminated sites belong to Sphingomonas (Pinyakong et al., 2000 2000 ), Paracoccus (Teng et al., 2010 2010 ), Mycobacterium (Luo, 2008 2008 ), Comamonas (Goyal and Zylstra, 1996 1996 ), Pseudomonas (Jacques et al., 2005 2005 ; Santos et al., 2008 2008 ), Burkholderia (Juhasz et al., 1997 1997 ), Janibacter (Zhang et al., 2009 2009 ), and Sphingobium (Cunliffe and Kertesz, 2006 2006 ). Enrichment of PAH-degraders is a common phenomenon during the course of PAH biodegradation (Guo et al., 1997 1997 ; Piskonen et al., 2005 2005 ). Therefore, in this study, the large enrichment of three TRFLP fragments might also be linked to the enrichment of some PAH-degraders. The greatly enriched fragments (217, 222, and 258 bp) ( Hae III digest) obtained from Sample B belong to Methylophilus , Mesorhizobium , and Terrimonas , respectively. However, surprisingly, to the best of our knowledge, no isolates from these genera have previously been founded in PAH-contaminated soils or linked to PAH biodegradation. Some species of Methylophilus have been linked to degradation or utilization of various compounds including dichloromethane (Nikolausz et al., 2006 2006 ), methylated amines (Large and Haywood, 1981 1981 ), methanol (Jenkins et al., 1987 1987 ), glucose, and galactose (Doronina and Trotsenko, 1994 1994 ). Two recent reports have also linked Methylophilus to utilization of biphenyl (Uhlik et al., 2009 2009 ), phenol, and humic matter (Hutalle-Schmelzer et al., 2010 2010 ). Interestingly, these compounds have a common phenolic hydroxyl proposed as intermediate compounds in anthracene degradation pathways (Pinyakong et al., 2003 2003 ). Therefore, Methylophilus may play roles in the degradation of intermediates of anthracene degradation, thus leading to its enrichment. Mesorhizobium species are usually isolated from root nodules (Lin et al., 2009 2009 ; Chen et al., 2010 2010 ). Recently, some species of Mesorhizobium have been linked to sulfur oxidization (Ghosh and Roy, 2006 2006 ), nitrile biotransformation (Feng et al., 2008 2008 ), oxamyl degradation (Osborn et al., 2010 2010 ), high arsenic resistance (Huang A.H. et al., 2010 2010 ), chromium reducing (Wani et al., 2009 2009 ), and acetonitrile biodegradation (Feng and Lee, 2009 2009 ). However, little is known about the environmental significances of Terrimonas . One recent report has revealed the dominance of Terrimonas in methyl tert-butyl ether degrading culture (Liu et al., 2009 2009 ). Therefore, the roles of Methylophilus , Mesorhizobium , and Terrimonas in PAH biodegradation still remained largely unclear. 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Microbial community changes in contaminated soils in response to phenanthrene amendment 10.1631/jzus.B1000440 Project (No. 50979002) supported by the National Natural Science Foundation of China 2010-12-20 2011-08-18 2011-03-08 2022-01-05 Copyright © Zhejiang University and Springer-Verlag Berlin Heidelberg 2011 2011 Journal of Zhejiang University-SCIENCE B (Biomedicine and Biotechnology) 09 12