Autochthonous Bioaugmentation-Modified Bacterial Diversity of

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Cite This: Environ. Sci. Technol. 2018, 52, 2934−2944

Autochthonous Bioaugmentation-Modified Bacterial Diversity of Phenanthrene Degraders in PAH-Contaminated Wastewater as Revealed by DNA-Stable Isotope Probing Jibing Li,†,¶ Chunling Luo,*,†,‡ Dayi Zhang,§ Mengke Song,‡ Xixi Cai,# Longfei Jiang,† and Gan Zhang† †

Guangzhou Institute of Geochemistry, Chinese Academy of Sciences, Guangzhou 510640, China College of Natural Resources and Environment, South China Agricultural University, Guangzhou, 510642, China § School of Environment, Tsinghua University, Beijing 100084, China # College of Resources and Environment, Fujian Agriculture and Forestry University, Fuzhou, 350002, China ¶ University of Chinese Academy of Sciences, Beijing, 100049, China ‡

S Supporting Information *

ABSTRACT: To reveal the mechanisms of autochthonous bioaugmentation (ABA) in wastewater contaminated with polycyclic aromatic hydrocarbons (PAHs), DNA-stable-isotopeprobing (SIP) was used in the present study with the addition of an autochthonous microorganism Acinetobacter tandoii LJ-5. We found LJ-5 inoculum produced a significant increase in phenanthrene (PHE) mineralization, but LJ-5 surprisingly did not participate in indigenous PHE degradation from the SIP results. The improvement of PHE biodegradation was not explained by the engagement of LJ-5 but attributed to the remarkably altered diversity of PHE degraders. Of the major PHE degraders present in ambient wastewater (Rhodoplanes sp., Mycobacterium sp., Xanthomonadaceae sp. and Enterobacteriaceae sp.), only Mycobacterium sp. and Enterobacteriaceae sp. remained functional in the presence of strain LJ-5, but five new taxa Bacillus, Paenibacillus, Ammoniphilus, Sporosarcina, and Hyphomicrobium were favored. Rhodoplanes, Ammoniphilus, Sporosarcina, and Hyphomicrobium were directly linked to, for the first time, indigenous PHE biodegradation. Sequences of functional PAH-RHDα genes from heavy fractions further proved the change in PHE degraders by identifying distinct PAH-ring hydroxylating dioxygenases between ambient degradation and ABA. Our findings indicate a new mechanism of ABA, provide new insights into the diversity of PHE-degrading communities, and suggest ABA as a promising in situ bioremediation strategy for PAH-contaminated wastewater.

1. INTRODUCTION Polycyclic aromatic hydrocarbons (PAHs) are ubiquitous environmental pollutants consisting of two or more fused aromatic rings.1 These compounds have the potential to pose serious health risks to all organisms.2 They are of great environmental concern due to their high toxicity, mutagenicity, and carcinogenicity.3 Phenanthrene (PHE) is employed as a model PAH compound due to its ubiquity in nature and fusedring angular structure.4 Bioremediation is considered as a costeffective and ecofriendly approach for elimination of PAHs from natural environments.5 PAHs in a polluted environment can be partially degraded by the autochthonous microbial population via natural attenuation. When indigenous microbes lack the capacity to degrade PAHs, the introduction of PAH degraders isolated from other PAH-contaminated sites,6,7 known as bioaugmentation (BA), is a promising technology to encourage PAH degradation.8,9 However, the efficacy of BA is variable, since © 2018 American Chemical Society

the survival and the PAH-degrading ability of the introduced microorganisms are highly dependent on environmental conditions.10,11 To overcome these difficulties, autochthonous bioaugmentation (ABA) has been proposed as an alternative method.12 ABA is defined as a BA technology that uses indigenous degraders at the polluted sites (soil, sediment, and water) with the aim of accelerating and enhancing the biodegradation potential.13 Many microorganisms capable of degrading PAHs have been isolated using cultivation-based approaches and have been evaluated for their potential application in ABA strategies.7,14,15 Cultivation-based methods are used to determine the Received: Revised: Accepted: Published: 2934

November 4, 2017 January 16, 2018 January 29, 2018 January 29, 2018 DOI: 10.1021/acs.est.7b05646 Environ. Sci. Technol. 2018, 52, 2934−2944

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Environmental Science & Technology

information regarding the mechanisms of ABA in aiding bioremediation of PAH-contaminated wastewater and other potentially polluted environments.

phenotypes and metabolic characteristics of PAH degraders and the functional genes associated with PAH metabolism such as PAH-ring hydroxylating dioxygenase.16 Particularly, the gene encoding the alpha subunit of the PAH-ring hydroxylating dioxygenase (PAH-RHDα) is usually used as the biomarker to quantify the bacterial populations capable of degrading PAHs.16 However, these approaches share an critical drawback in that the majority of microbes are uncultivable,17 and many autochthonous PAH-degrading microorganisms remain resistant to traditional cultivation approaches.4,18 In addition, direct cultivation greatly underestimates the microbial diversity in ecosystems and fails to explain the complex interactions between microorganisms and the natural environment.19 In recent years, cultivation-independent methods have been used to evaluate the prokaryotic diversity of complex systems and estimate the PAH degradation potential of indigenous consortia.20−22 High-throughput sequencing has revolutionized our ability to investigate the microbial communities in environmental samples by providing higher resolution of microbial taxa compared with conventional cloning techniques.23 Nevertheless, the metabolic features of organisms are difficult to infer accurately using this approach.23 Stable-isotope probing (SIP) is a cultivation-independent technique that circumvents the requirement of isolating an organism to assess metabolic responses and link its identity to function.24 This technique relies on the incorporation of a stable isotope-labeled substrate and the identification of active microorganisms by analyzing isotopeenriched cellular components, such as DNA, RNA, and protein.4,25 We can therefore characterize the target microorganisms resistant to cultivation.18 To date, SIP has been used to identify a large number of indigenous bacteria capable of degrading PAHs.20,26−28 In our previous work, we isolated a novel PHE degrading strain, Acinetobacter tandoii LJ-5, from PAH-contaminated wastewater (37°49′N, 118°25′E; altitude, 37.49 m) collected in Shandong Province, China in September 2015.18 A. tandoii LJ5 is a major autochthonous PHE degrader due to its PHE metabolism confirmed by its enrichment in the heavy DNA fraction in the presence of 13C-PHE.18 A. tandoii LJ-5 is a Gram negative, rod-shaped, obligate aerobe lacking flagella. This strain grew well in minimal medium supplemented with 100−1000 mg/L PHE under optimal conditions (pH 7.0 and 30 °C), and a satisfactory PHE degradation efficiency (>80%) was achieved within 3 days for low PHE concentrations (100 and 200 mg/ L).18 Although the ABA strategy can significantly promote PHE mineralization by introducing autochthonous strains, altering the overall microbial activity and population,15,29 no studies have confirmed the degrading capacity of the reintroduced degraders, and little is known about changes in the diversity of indigenous PHE degraders during the ABA. In the present study, wastewater microcosms were established to evaluate the bioremediation potential of ABA in PAH-contaminated wastewater with A. tandoii LJ-5 inoculum. Additionally, DNA-SIP was used to investigate the changes in the diversity of indigenous PHE degraders by comparing treatments in the presence versus absence of LJ-5, and to link the indigenous bacterial taxa with their PHE biodegradation phenotypes. The PHE-degrading bacteria in different wastewater treatments were successfully characterized using DNA-SIP and high-throughput sequencing. Furthermore, the functional PAH-RHDα gene was investigated by analyzing relevant sequences amplified from the heavy fractions of 13C-PHE treatments. We have provided useful

2. MATERIALS AND METHODS 2.1. Sample Collection. Water samples (∼10 L) were collected in March 2016 from industrial wastewater at the same oil refinery (37°49′N, 118°25′E; altitude, 37.49 m) in Shandong Province (China) as described in our previous work.18 After transport to the laboratory, the wastewater was stored immediately at 4 °C. Portions of the samples were stored at −80 °C for subsequent DNA extraction, and the remainder was promptly used in PHE degradation and SIP experiments. PAHs identified in the wastewater are listed in Table S1 (determined using gas chromatography−mass spectrometry as described below). 2.2. Cultivation of Bacteria. A. tandoii LJ-5 was previously isolated from wastewater in September 2015 at the same oil refinery, using PHE as the carbon source,18 and deposited in the Marine Culture Collection of China (MCCC; 1K02142). A. tandoii LJ-5 is a promising candidate for the remediation of PAHcontaminated water, as it has been shown to be an indigenous PHE degrader by both DNA-SIP and direct cultivation. A. tandoii LJ-5 was stored in glycerol solution (15% v/v) at −80 °C, thawed, and recultivated as described previously.18 Briefly, it was cultivated in minimal medium supplemented with 1000 mg/L PHE in the dark for 18−24 h with shaking at 180 rpm at 30 °C. Enriched cultures were transferred into a sterile centrifuge tube. Pelleted cells were obtained after centrifugation for 10 min (3000 × g; Eppendorf, Centrifuge 5810R, Hamburg, Germany) and resuspended in phosphate-buffered saline (PBS). Finally, the population of A. tandoii LJ-5 was determined to be approximately 2 × 108 colony forming units/mL (CFU/mL), using the dilution plate counting method.30 Details of the surface tension measurement were provided in the Supporting Information. 2.3. SIP Microcosms. Microcosms were established in 150 mL serum bottles with 50 mL wastewater with and without inoculation of strain LJ-5. For treatments with strain LJ-5, 100 μL of the cell/PBS suspension were added to the wastewater, thereby establishing an initial inoculum population of approximately 4 × 105 CFU/mL at the beginning of each treatment.10 The unlabeled PHE (99%) or 13C-labeled PHE (13C14−PHE, 99%, Cambridge Isotope Laboratories, Inc., Tewksbury, MA, USA) was added to the above bottles at a final concentration of 10 mg/L. The biotic treatments, which allowed comparisons of ambient degradation (AD) and ABA, are referred to as 12C_AD (12C-PHE without strain LJ-5), 12C_ABA (12C-PHE with strain LJ-5), 13C_AD (13C-PHE without strain LJ-5), and 13C_ABA (13C-PHE with strain LJ-5). Microcosms without PHE (nonPHE control) and treatments with unlabeled PHE in filtersterilized wastewater (sterile control) were also established. Each treatment was performed in triplicate. The microcosms were incubated using the method described in our previous work.18 The residual PHE on days 3 and 6 after incubation was extracted and detected. Almost all of the PHE was degraded on day 6 (Figure S3), indicating that the majority of PHE was mineralized and incorporated by the microbes during the first 3 days. Thus, samples from each treatment were collected on day 3 for PHE analysis and DNA extraction to avoid cross-feeding. 2.4. Nucleic Acid Extraction and Ultracentrifugation. DNA extraction from triplicate microcosms (25 mL) of each treatment was performed and quantified as described previously.18 Ultracentrifugation of DNA was conducted as 2935

DOI: 10.1021/acs.est.7b05646 Environ. Sci. Technol. 2018, 52, 2934−2944

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Environmental Science & Technology described in our previous report.31 Briefly, approximately 3 μg DNA from each microcosm were added to Quick-Seal polyallomer tubes (13 × 51 mm, 5.1 mL, Beckman Coulter, Pasadena, CA, USA) and mixed with Tris-EDTA (pH 8.0)/ cesium chloride solution at a final buoyant density (BD) of ∼1.77 g/mL. After balancing and heat-sealing, density gradient centrifugation was performed in an ultracentrifuge (Optima L100XP, Beckman Coulter, USA) at 45,000 rpm (20 °C) for 48 h. Centrifuged gradients were fractionated into different fractions of 400 μL each. The DNA fractions were purified after the BD of each fraction was measured. The relationships between BD and the fraction number or DNA concentration for all four biotic treatments are listed in Figures S1 and S2, respectively. 2.5. Quantitative Polymerase Chain Reaction (qPCR). Abundance of the bacterial 16S rRNA gene in each fraction was determined by qPCR using the universal bacterial primer pair Bac519F/Bac907R (Table S2). The 20 μL qPCR mixture contained 10 μL of SYBR green PCR Premix Ex TaqII (Takara Bio Inc., Japan), 0.2 μL of each primer (10 μM), and 1 μL of DNA template. The qPCR was performed using the ABI 7500 real-time PCR system (Applied Biosciences, USA). Ten-fold serial dilutions of known copy numbers of the plasmid DNA extracted from Escherichia coli were generated to produce a standard curve. The reactions were conducted as follows: initial denaturation at 94 °C for 10 min, followed by 40 cycles at 94 °C for 30 s, 55 °C for 30 s, and 72 °C for 15 s. The SYBR green signal was measured after a 20 s step at 72 °C in each cycle. The melt curve was obtained by heating from 60 to 95 °C at a rate of 0.2 °C/s. 2.6. High-Throughput Sequencing and Computational Analyses. The hypervariable V4 region of bacterial 16S rRNA gene fragments from the DNA fractions derived from the 12 C_AD, 12C_ABA, 13C_AD, and 13C_ABA microcosms was amplified using the 515f/806r primer set (Table S2), as described previously.32 Unique heptad-nucleotide sequences (12 bases) were added to the reverse primers as barcodes to assign sequences to the different fractions. PCR was performed using the method described by Song et al.33 Sequencing was conducted using 2 × 250 base pair paired-end technology on the Illumina MiSeq sequencer in a standard pipeline. The qualified sequences were analyzed after read filtering34,35 and assigned using an operational taxonomic unit (OTU)-based method to generate microbiome profiles, as described previously.36−38 OTU assignment was based on a 97% cutoff in the present study. The relative abundance of each OTU was determined, and the top 100 most relatively abundant ones were selected for analysis according to previous studies.18,39 The PHE degraders were identified by OTUs enriched in the heavy fractions from the 13 C_AD and 13C_ABA samples compared with the 12C_AD and 12 C_ABA samples. In the present study, we identified four OTUs (OTU_56, OTU_65, OTU_106, and OTU_111) from the 13 C_AD treatment and seven OTUs (OTU_8, OTU_15, OTU_23, OTU_43, OTU_83, OTU_106, and OTU_111) from the 13C_ABA treatment. These sequences were further trimmed using the Greengenes database and aligned to Bacillus spp. (OTU_8, accession number MF037428), Ammoniphilus spp. (OTU_15, accession number MF037425), Hyphomicrobium spp. (OTU_23, accession number MF037426), Paenibacillus spp. (OTU_43, accession number MF037429), Rhodoplanes spp. (OTU_56, accession number MF037430), Xanthomonadaceae spp. (OTU_65, accession number MF037431), Sporosarcina spp. (OTU_83, accession number

MF037427), Mycobacterium spp. (OTU_106, accession number MF037432), and Enterobacteriaceae spp. (OTU_111, accession number MF037433), respectively. Phylogenetic analysis of these sequences was performed as described previously.4 2.7. Detection of PAH-RHDα Genes. The PAH-RHDα genes in the heavy DNA fractions of the 13C-PHE treatment samples were amplified using two primer sets for Gram positive (GP; 642f/933r) and Gram negative (610f/911r) degraders,33 respectively (Table S2). Gradient PCR and the amplification reactions were performed as described previously.16 However, in the present study, only the PAH-RHDα GP primer set produced a strong and specific amplicon and was selected for analysis. We identified two PAH-RHDα GP genes (PAH-RHDα C1 and C15) from the 13C_AD treatment, and three PAH-RHDα GP genes (PAH-RHDα L1, L6, and C10) from the 13C_ABA treatment. These sequences are available in GenBank with the following accession numbers: PAH-RHDα C1, MF037434; PAH-RHDα C15, MF037435; PAH-RHDα L1, MF037436; PAH-RHDα L6, MF037437; PAH-RHDα L10, MF037438. 2.8. Chemical Analysis. The concentrations of PHE and other PAHs in each microcosm were analyzed by gas chromatography (model 7890, Agilent, Santa Clara, CA, USA) using a capillary column (DB-5MS, 30 m, 0.25 mm, 0.25 μm) and a mass spectrometric detector (model 5975, Agilent) on days 0, 3 and 6, as described previously.18 Briefly, the water sample was spiked with 1,000 ng deuterated PAHs and extracted twice with dichloromethane. The extracted organic phase was concentrated to approximately 0.5 mL and purified using a silica gel/alumina column (8 mm i.d.). The eluent was concentrated to approximately 50 μL using a gentle stream of N2, and 1000 ng of hexamethylbenzene were added as an internal standard to all samples before instrumental analysis. The components and concentrations of the deuterated PAHs, standards and the internal standard are listed in Table S3.

3. RESULTS 3.1. PHE Degradation. The PHE biodegradation curves in the 12C_AD, 12C_ABA, 13C_AD, and 13C_ABA treatments are shown in Figure S3. The recovery rates of PHE during the extraction procedure were 79−85% in this study. The PHE concentration in the sterile control exhibited less decrease than those in the biotic treatments, consistent with our previous observations. On day 3, there was 36.7%, 38.2%, 22.6%, and 23.8% residual PHE in the 12C_AD, 13C_AD, 12C_ABA, and 13 C_ABA microcosms, respectively, indicating that PHE biodegradation occurred in the biotic treatments. No significant difference (p > 0.05) was observed between either the 12C_AD and 13C_AD treatments, or the 12C_ABA and 13C_ABA treatments. Moreover, the PHE concentrations in the 12 C_ABA (22.6%) and 13C_ABA (23.8%) microcosms were significantly lower (p < 0.05) than those in the 12C_AD (36.7%) and 13 C_AD (38.2%) microcosms, suggesting that the degradation efficiency increased by 13.8% when strain LJ-5 was added to the wastewater. Statistical analysis showed a significant difference between the AD and ABA treatments after 3 days of incubation, suggesting that ABA strategy by A. tandoii LJ-5 inoculum could significantly improve the PHE biodegradation efficiency in PAH-contaminated wastewater. In addition, there was little loss of other PAHs in the wastewater over the 3-day period (data not shown). 3.2. PHE Degraders Revealed by DNA-SIP. DNA extracted from all four biotic treatments, after 3 days of 2936


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Figure 1. Shift tendency of OTU_56, OTU_65, OTU_111, and OTU_106 fragments. The relative abundances of the OTU_56, OTU_65, OTU_111, and OTU_106 fragments over a range of buoyant density (BD) of DNA extracted from the 12C_AD and 13C_AD treatments. Mean ± standard deviation (SD) relative abundance from triplicate microcosms are shown.

Figure 2. Shift tendency of OTU_8, OTU_15, OTU_43, OTU_83, OTU_23, OTU_106, and OTU_111 fragments. The relative abundance of the OTU_8, OTU_15, OTU_43, OTU_83, OTU_23, OTU_106, and OTU_111 fragments over a range of buoyant density (BD) of DNA extracted from the 12C_ABA and 13C_ABA treatments. Mean ± standard deviation (SD) relative abundance from triplicate microcosms are shown. 12

C_AD and 13C_AD treatments or the 12C_ABA and 13C_ABA treatments (Figure S4). Moreover, the addition of strain LJ-5 induced a significant change in the composition and structure of microbial communities in the ABA treatments, compared with the AD microcosms. Obviously, the relative abundance of the

incubation, were separated by isopycnic cesium chloride density gradient centrifugation, followed by high-throughput sequencing of each fraction. The relative abundance of the total 16S rRNA defined by the genus indicated a slight difference in indigenous microbial communities between the samples from either the 2937


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Figure 3. Phylogenetic tree of identified OTUs responsible for PHE degradation. Neighbor-joining tree based on 16S rRNA gene sequences showing the phylogenetic position of the bacteria corresponding to OTU_8, OTU_15, OTU_23, OTU_43, OTU_83, OTU_56, OTU_65, OTU_111, OTU_106, and their representatives of other related taxa. Bootstrap values (expressed as percentages of 1000 replications) > 50% are shown at the branch points. Bar 0.05 substitutions per nucleotide position.

unclassified Xanthomonadaceae, were enriched in the AD microcosms (Figure S4). The relative abundance of the genus Kaistobacter (38.4%) increased and was significantly higher (p < 0.05) in the ABA microcosms than those in the AD microcosms (22.9%), whereas the relative abundances of the genera

genus Acinetobacter (2.08%) increased significantly after addition of strain LJ-5 in the ABA treatments, and was much higher than that in the AD microcosms (0.05%). The majority of the abundant bacteria (>5% total abundance), including members of the genera Kaistobacter, Rhodanobacter, Burkholderia, and 2938


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Figure 4. Phylogenetic tree of amplified PAH-RHDα GP genes from the heavy fractions of the 13C_AD and 13C_ABA microcosms. PAH-RHDα C1 and C15 represent the PAH-RHDα genes amplified from the heavy fractions of the 13C_AD treatment, and PAH-RHDα L1, L6, and L10 represent the PAHRHDα genes from the 13C_ABA treatment.

Rhodanobacter, Burkholderia, and unclassified Xanthomonadaceae were significantly lower (p < 0.05) in the ABA microcosms (15.8%, 1.9% and 3.3%, respectively) than those in the AD microcosms (24.3%, 7.2% and 10.5%, respectively). As for rare bacteria, the ABA strategy stimulated growth of the genera Conexibacter and Rhodoplanes, owing to their higher relative abundances (2.09% and 2.01%) compared with the AD microcosms (0.76% and 1.23%, p < 0.05). No significant changes were detected in the non-PHE controls (Figure S5). The bacterial 16S rRNA gene abundance was quantified by qPCR using DNA recovered from each fraction of all samples as the template. As shown in Figure S6, in both AD and ABA microcosms after the 3-day incubation, the bacterial 16S rRNA in fractions with higher BDs (1.7382 or 1.7556 g/mL; see asterisks in Figures 1 and 2) was significantly higher in the 13C-PHE treatments than those in the 12C-PHE control (marked in gray). The abundance of total 16S rRNA of additive bioaugmented strain LJ-5 was 4 × 105 copies/mL, accounting for 2% of that in the ABA treatment (1.96 × 107 copies/mL; Figure S7). Accordingly, the relative abundance of the genus Acinetobacter after ABA treatment increased to 2.08%, and the total 16S rRNA gene copy number of ABA microcosms was slightly higher than that of raw wastewater (Figure S7), indicating that the bioaugmented strain LJ-5 persisted in these treatments. The indigenous microorganisms responsible for 13C-PHE assimilation were detected by comparing the relative abundances of specific OTUs in the 12C-PHE and 13C-PHE treatments from

each fraction. As shown in Figure 1, OTU_56, OTU_65, OTU_106, and OTU_111 were enriched at higher BDs (1.7382 or 1.7556 g/mL) in the 13C_AD microcosms but not in the 12 C_AD treatments. Comparing to the relative abundances of OTU_56, OTU_65, OTU_106, and OTU_111 in the same fractions of the 12C_AD treatment (0.18%, 0.15%, 0.11%, and 0.82%, respectively), the higher abundance in the heavy fractions from the 13C_AD sample (1.49%, 1.68%, 0.94%, and 4.45%, respectively) indicated that the microorganisms represented by the above OTUs played a primary role in PHE degradation. However, the inoculation of strain LJ-5 produced a significant change in the diversity of the indigenous PHE-degrading communities. Seven main types of bacteria represented by OTU_8, OTU_15, OTU_23, OTU_43, OTU_83, OTU_106, and OTU_111 at a higher BD (1.7382 or 1.7556 g/mL) were enriched in the 13C_ABA sample, but not in the 12C_ABA sample (Figure 2). Similarly, the relative abundances of OTU_8, OTU_15, OTU_23, OTU_43, OTU_83, OTU_106, and OTU_111 were significantly higher in the heavy fractions of the 13C_ABA treatment (0.93%, 1.71%, 1.26%, 1.08%, 0.11%, 0.31%, and 7.78%, respectively) than those in the 12C_ABA microcosm (0.05%, 0.08%, 0.01%, 0.01%, 0.007%, 0.13%, and 1.55%, respectively). Nevertheless, OTU_4, representing strain LJ-5, was not enriched in the heavy fractions of the 13C_ABA treatment samples because of its lower relative abundance in the heavy fractions (higher BDs, 1.7382 and 1.7556 g/mL) of the 13 C_ABA microcosm (0.02 and 0.02%, respectively) than in the 2939


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rium vanbaalenii PYR-1 (AAY85176.1), Mycobacterium novocastrense (GAT12202.1), and Terrabacter sp. FLO (ABA87073.1), and PAH-RHDα L10 shows 100% similarity with the PAH-RHD genes of Bacillus sp. CL1-1 (AND66078.1), Enterobacter sp. CL12 (AND66079.1), and uncultured Comamonas sp. (CAN85224.1), and forms a subclade with a bootstrap value of 61.

C_ABA microcosm (0.13 and 0.05%, respectively). Although PHE-degrading communities differed between the AD and ABA treatments, OTU_106 and OTU_111 played a vital role in PHE degradation under both conditions. Figure 3 shows the phylogenetic information for the PHE degraders represented by the above OTUs. OTU_8, OTU_15, OTU_43 and OTU_83 belong to the genus Bacillus (family Bacillaceae), Ammoniphilus (family Aneurinibacillus), Paenibacillus (family Paenibacillaceae) and Sporosarcina (family Planococcaceae), respectively, within the same order Bacillales (phylum Firmicutes, class Bacilli). OTU_15 shares 100% similarity with Ammoniphilus resinae CC-RT-E (HM193518) and forms a subclade with a high bootstrap value of 99. OTU_43 exhibits 98.02% similarity with the partial 16S rRNA gene sequence of strain Paenibacillus xylanisolvens X11-1 (AB495094). OTU_83 has 100% similarity with strains Sporosarcina ureae DSM 2281 (AF202057), Sporosarcina aquimarina SW28 (AF202056), Sporosarcina saromensis HG645 (AB243859), and Sporosarcina luteola Y1 (AB473560) and forms a subclade with a bootstrap value of 84. OTU_23 and OTU_56 are assigned to the genus Hyphomicrobium (family Hyphomicrobiaceae) and Rhodoplanes (family Bradyrhizobiaceae), respectively, within the same order Rhizobiales (phylum Proteobacteria, class Alphaproteobacteria). OTU_23 exhibits 98.81% similarity with the partial 16S rRNA gene sequence of strain Hyphomicrobium denitrif icans ATCC 51888 (ACVL01000012), and forms a subclade with a high bootstrap value of 96. OTU_56 shares 100% similarity to the partial 16S rRNA gene sequence of uncultured Rhodoplanes sp. clone Leob163 (KF226093.1), and forms a subclade with a high bootstrap value of 98. OTU_65 and OTU_111 are classified in the family Xanthomonadaceae (order Xanthomonadales) and Enterobacteriaceae (order Enterobacteriales), respectively, within the same class Betaproteobacteria (phylum Proteobacteria). Additionally, OTU_106 is characterized as genus Mycobacterium (phylum Actinobacteria, class Actinobacteria, order Corynebacteriales, family Mycobacteriaceae) and shares 100% similarity with many strains in this genus, including Mycobacterium brisbanense ATCC 49938 (AY012577), Mycobacterium elephantis 484 (AJ010747), Mycobacterium moriokaense DSM 44221 (AJ429044), and Mycobacterium pulveris DSM 44222 (AJ429046), and forms a subclade with a high bootstrap value of 100. 3.3. Presence of PAH-RHDα Genes Involved in PHE Metabolism. In the present study, the PAH-RHDα GP genes were analyzed in the heavy fractions of the 13C_AD and 13 C_ABA treatments (marked with an asterisk in Figures 1 and 2). PAH-RHDα C1 and C15 were detected in the heavy fractions of the 13C_AD treatment (Figure 4). The 13C_ABA treatment produced a conspicuous change in the types of PAH-RHDα GP genes, compared with the AD microcosms, and PAH-RHDα L1, L6, and L10 were detected (Figure 4). Thereinto, PAH-RHDα C1 and L1 share 100% similarity and are the same PAH-RHDα GP gene, indicating that they exist in the heavy fractions of both 13 C_AD and 13C_ABA microcosms. They form a subclade with a high bootstrap value of 96 in the phylogenetic tree of amplified PAH-RHDα GP genes, and show 89% similarity with the PAHRHD genes of Mycobacterium sp. S23 (ALS30442.1). The other PAH-RHDα gene in the heavy fractions of the 13C_AD treatment, PAH-RHDα C15, exhibits 99% similarity with the PAH-RHD genes of uncultured bacterium (AMM73080.1). In the heavy fractions of the 13C_ABA treatment, PAH-RHDα L6 reveals 99% similarity with the PAH-RHD genes of Mycobacte-

4. DISCUSSION DNA-SIP has been used successfully to demonstrate that native microorganisms collected at field sites are involved in PHE biodegradation.20,27 Gutierrez et al. used DNA-SIP with fully 13 C-labeled PHE to link the phylogenetic identity of bacterial taxa with their ability to mineralize PHE in surface slicks and plume waters.27 Jones et al. performed DNA-SIP with 13Clabeled PHE as part of a larger project investigating strategies for bioremediation of PAH-contaminated soil from a former manufactured-gas plant site.20 Our work applied 13C-PHE as the substrate in DNA-SIP and revealed the mechanisms of ABA strategy in PAH-contaminated wastewater with the addition of the autochthonous microorganism A. tandoii LJ-5. Although ABA strategies have been successfully applied for the remediation of PAH-contaminated sites,14,15 limited studies have addressed the changes that occur in the microbial community during the ABA process.14,40 In the present study, ABA with strain LJ-5 inoculum produced a significant increase in PHE biodegradation efficiency in PAH-contaminated wastewater. This suggests that strain LJ-5 is potentially an ABA agent encouraging the remediation of PAH-contaminated sites. Moreover, the addition of LJ-5 markedly modified the bacterial community structure. By comparing the structures of microbial communities between AD and ABA, we identified three bacterial genera, affiliated with Kaistobacter, Conexibacter and Rhodoplanes, that were stimulated by LJ-5 inoculation. Kaistobacter demonstrated relatively high abundance in the indigenous microbial communities of a tailings dump contaminated with antimony41 and the ability to biodegrade both S-ethyldipropylthiocarbamate and atrazine in soils.42 Members of the genus Conexibacter within the class Actinobacteria are affiliated with aromatic hydrocarbon degraders and contain bph genes encoding the biphenyl degradative pathway.43 Rhodoplanes has been proposed to accommodate taxa that are primarily phototrophic and dominant in rhizospheric soils and roots.44 Furthermore, the addition of LJ5 also affected the composition of abundant bacteria, including members of the genera Kaistobacter, Rhodanobacter, Burkholderia, and unclassified Xanthomonadaceae. Among them, Rhodanobacter spp. are dominant in crude oil plus dispersant and rhizospheres of maize cultivars,45,46 and Burkholderia and unclassified phylotypes within the family Xanthomonadaceae, such as Stenotrophomonas, are the dominant genera of PAH degraders with high metabolic activities in PAH-contaminated soils and sediments.47,48 However, there was no previous evidence directly linking the above activated or abundant bacteria (except Xanthomonadaceae and Burkholderia) to PHE degradation. They might play important roles in stabilizing the microbial community or degrading other organic pollutants in PAH-contaminated wastewater.18 Besides the structure and dynamics of the microbial community, ABA might also shape the composition of functional PHE-degrading community, although this has not yet been reported. Our study is the first to show that ABA also influences the abundance and diversity of PAH-degrading bacteria in PAHpolluted wastewater, suggesting that different PHE degradation 2940


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obligately oxalotrophic, and motile by peritrichous flagella.64 Strains of the genus Ammoniphilus are able to utilize high concentrations of ammonium ions.65 Similar to Ammoniphilus, Sporosarcina also belongs to the phylum Firmicutes. It is usually detected in clinical specimens and raw cow’s milk.66 Sporosarcina has been shown to principally degrade aflatoxin B-1, which is the most potent naturally occurring carcinogen known.67 Hyphomicrobium, belonging to the Alphaproteobacteria and Hyphomicrobiaceae family, has versatile metabolic capabilities such as the ability to degrade dichloromethane,68 methamidophos,69 dimethyl sulfide, and methanol.70 However, the taxa Ammoniphilus, Sporosarcina, and Hyphomicrobium have not been linked previously with PHE degradation; thus, their mechanisms of degradation in PAH-contaminated sites remain unknown. Our results provide strong evidence that members of the above genera are the primary indigenous PHE degraders induced by ABA strategy in wastewater, expanding our knowledge of the diversity of indigenous PHE-degrading communities. Despite the satisfactory performance of ABA strategies in bioremediation by the addition of autochthonous degraders, no studies have confirmed the activities or degrading capacity of the reintroduced strains. It is worth mentioning that, although the ABA strategy with LJ-5 inoculum encouraged PHE biodegradation, the participation of A. tandoii LJ-5 in PHE degradation in situ was questioned due to their limited enrichment in the heavy DNA fraction according to the DNA-SIP results. Previous researchers have focused primarily on the effects of reintroduced degraders on degradation efficiency and the total microbial activities or population during ABA.15,29 The diversity and functions of the whole microbial community can be enhanced by adding a strain to the polluted sites.40,71 Our work is the first study to uncover the roles of the autochthonous strain LJ-5 in ABA strategy, modifying the diversity of indigenous PHE degraders instead of participating in PAH degradation in situ. Such a mechanism has never been addressed before and remains unclear in PAH-contaminated wastewater. One possible reason might be attributed to the change in wastewater composition, consequently leaving strain LJ-5 from functional PHE-degraders in previous wastewater to inactive microbes in the present study. Another reason might be the contribution of Acinetobacter biosurfactants to the bioremediation of PAHs and the shift in community composition, which has been confirmed by other researchers.72 For instance, a wide range of Acinetobacter sp. can produce biosurfactants (amphiphilic molecules consisting of hydrophilic and hydrophobic domains),73 which might increase the bioavailability of hydrophobic compounds by solubilization and/or emulsification or alter the cell surface properties of other microorganisms.72 In the present study, A. tandoii LJ-5 could reduce the surface tension of water from 70.0 to 35.2 mN/m. Our results indicated that strain LJ-5 might be capable of producing biosurfactants, simultaneously improving PHE bioavailability and activating the functions of other PHE-degrading candidates. Moreover, the relative abundance of the bioaugmented strain LJ5 was 4 × 105 copies/mL, accounting for 2% of the total 16S rRNA in ABA treatments and approximately 2 times as that of the initial inoculum abundance in surface tension measurements. It suggested sufficient amount of strain LJ-5 in all the bioaugmentation treatments to enhance PHE degradation via the production of biosurfactants. More evidence of a change in the PHE-degrading community was provided by the sequences of PAH-RHDα genes between AD and ABA treatments. Only one novel PAH-RHDα GP gene, PAH-RHDα C1 and L1, was detected in the heavy DNA fraction

activities can be achieved by distinct microbial communities within the same environment. In the AD treatment, the indigenous microorganisms responsible for PHE degradation were affiliated with Rhodoplanes, Mycobacterium, Xanthomonadaceae (genus unclassified), and Enterobacteriaceae (genus unclassified). The genus Rhodoplanes classified under the family Hyphomicrobiaceae of the order Rhizobiales in the class Alphaproteobacteria, was first described in 1994.49 Members of this genus are characterized as Gram negative, phototrophic, rodshaped, motile and are widespread in aquatic habitats.50 Rhodoplanes is the dominant genus in wastewater treatment plants,51 and some strains in this genus can grow on organic carbon sources, such as acetate, pyruvate, glucose, and fructose.50 However, there is still no proof of its ability to degrade PAHs, including PHE, and their roles in PAH-contaminated sites remain unclear. Our results provide unequivocal evidence that some microbes in this genus are primarily responsible for PHE degradation in PAH-polluted wastewater. Mycobacterium is wellknown to degrade various environmental contaminants such as chlorinated compounds, polychlorobiphenyls, and PAHs.52 Degradation of PHE by Mycobacterium has been reported previously,53 and several strains in this genus can effectively degrade high molecular weight PAHs.54 However, prior to this study, PHE degradation by indigenous Mycobacterium has not been identified using DNA-SIP. The families Xanthomonadaceae and Enterobacteriaceae are members of the class Betaproteobacteria. Previous studies have reported that Stenotrophomonas sp. and Pseudoxanthomonas sp. in the family Xanthomonadaceae possess the functions of metabolizing a wide range of PAHs, for example, naphthalene, PHE, anthracene, fluorene, pyrene, and benzo[a]pyrene.55−57 Members of the family Enterobacteriaceae, such as Raoultella sp. and Klebsiella sp., can also degrade PAHs efficiently, including acenaphthene, fluorene, pyrene, benzo[a]pyrene, PHE, and fluoranthene.58−60 Here, our results provide strong evidence that some microbes in these families are active PHE degraders in PAH-contaminated wastewater. It should be noted that the bacterial community structure of the wastewater in this study was significantly different from that of our previously reported wastewater,18 thus resulting in the change in diversity of microorganisms responsible for in situ PHE degradation. The majority of the abundant bacteria in wastewater collected in September 2015 were affiliated with the genera Pseudomonas (23.5 ± 1.2%), unclassified Chitinophagaceae (4.7 ± 0.3%), and Comamonadaceae (6.4 ± 0.3%).18 We identified four OTUs directly responsible for indigenous PHE biodegradation, including the phylotypes affiliated with Acinetobacter, Sphingobium, Kouleothrix, and Sandaracinobacter. In the present study, among the PHE degraders identified in the AD microcosms, only two taxa Mycobacterium and Enterobacteriaceae (genus unclassified) were also identified as PHE degraders in the ABA treatments with strain LJ-5. Five new indigenous PHE-degrading organisms became active, namely phylotypes affiliated with Bacillus, Paenibacillus, Ammoniphilus, Sporosarcina, and Hyphomicrobium. The genera Bacillus and Paenibacillus are known to be metabolically versatile, degrading aromatic and hydroxylated aromatic compounds.55,61,62 Moreover, these two genera can degrade PHE,55 and appear to be highly competitive by the addition of root exudates.63 However, no studies have considered the roles of Bacillus and Paenibacillus in the remediation of PAH-contaminated sites by ABA strategy. The genus Ammoniphilus is a member of the family Aneurinibacillus in the phylum Firmicutes. Members of this genus are typically Gram variable, oxidase- and catalase-positive, 2941


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Environmental Science & Technology of both the 13C_AD and 13C_ABA treatments, possibly associated with Mycobacterium (OTU_106) or Xanthomonadaceae (genus unclassified, OTU_111), which were also both detected as functional PHE-degrading bacteria in the two microcosms. In addition, the distinctive PAH-RHDα L15 gene is potentially the functional gene from Rhodoplanes or Enterobacteriaceae (genus unclassified) as they were all detected in the heavy DNA fraction of the 13C_AD treatment. Accordingly, in the 13C_ABA microcosm, PAH-RHDα L6 and L10 genes might be linked to the PHE degraders of Bacillus, Paenibacillus, Ammoniphilus, Sporosarcina, or Hyphomicrobium. Here, the active PHE degraders might possess other functional genes that were not targeted by the primers used in this study, and we could not accurately attribute one functional gene to one bacterium for the lack of available database information. Indeed, little information is available regarding the functional genes of certain dominant PHE-degrading genera, such as Rhodoplanes, Ammoniphilus, Sporosarcina, and Hyphomicrobium, since this is first report of PHE biodegradation for these genera. ABA by A. tandoii LJ-5 was first applied as a potential strategy to enhance the remediation of PAH-contaminated wastewater in this study. Besides resulting in a significant increase in the PHE biodegradation efficiency, ABA remarkably modified the functional PHE degrading community of the wastewater. The indigenous microorganisms responsible for PHE degradation were affiliated with Rhodoplanes, Mycobacterium, Xanthomonadaceae and Enterobacteriaceae in the AD treatments, whereas five new taxa (Bacillus, Paenibacillus, Ammoniphilus, Sporosarcina, and Hyphomicrobium) were activated in PHE biodegradation in situ in the ABA microcosms. Of all the above PHE degraders, Rhodoplanes, Ammoniphilus, Sporosarcina, and Hyphomicrobium were linked to indigenous PHE biodegradation for the first time. Nevertheless, LJ-5 did not participate in indigenous PHE degradation. The change in PAH-RHDα gene diversity further confirmed our findings by different PAH-RHDα genes involved in PHE metabolism in the AD and ABA treatments. Collectively, our findings raise a new mechanism of ABA, provide new insights into the diversity of PHE-degrading communities and suggest ABA as a promising in situ strategy for PAH-contaminated wastewater.

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treatments; the 16S rRNA gene copies in the 12C_AD, C_AD, 12C_ABA, and 13C_ABA microcosms (PDF)

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AUTHOR INFORMATION

Corresponding Author

*E-mail: [email protected]. Tel.: +86-20-85290290. Fax: +86-2085290706. ORCID

Jibing Li: 0000-0003-3115-9017 Chunling Luo: 0000-0003-2359-4246 Dayi Zhang: 0000-0002-4175-5982 Gan Zhang: 0000-0002-9010-8140 Notes

The authors declare no competing financial interest.

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ACKNOWLEDGMENTS Financial support was provided by the Scientific and Technological Planning Project of Guangzhou, China (Nos. 201707020034), the National Natural Science Foundation of China (No. 41673111), and the Department of Science and Technology of Guangdong province (2016TQ03Z938).

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REFERENCES

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ASSOCIATED CONTENT

S Supporting Information *

The Supporting Information is available free of charge on the ACS Publications website at DOI: 10.1021/acs.est.7b05646. Concentrations of PAHs in wastewater; primers used for the PCR of 16S rRNA and PAH-RHD gene; the components of deuterated PAHs, standards, and internal standard; correlation between the fraction number and buoyant density of DNA extracted from the 12C-AD and 13 C-AD treatments and from the 12C-ABA and 13C-ABA treatments; correlation between DNA concentration and buoyant density from DNA extracted from (a) the 12C-AD and 13C-AD treatments, and (b) the 12C-ABA and 13CABA treatments; residual PHE percentage in wastewater after 3 days of incubation; relative abundance of 16S rRNA defined taxa by genus in the 12C_AD, 13C_AD, 12C_ABA, and 13C_ABA microcosms; relative abundance of 16S rRNA defined taxa by genus in non-PHE microcosms; the quantitative distribution of density-resolved bacterial 16S rRNA obtained from wastewater samples in the 12C-AD and 13C-AD treatments and the 12C-ABA and 13C-ABA 2942


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