Microbiome changes and oxidative capability of an anaerobic PCB dechlorinating enrichment culture after oxygen exposure

New BIOTECHNOLOGY 56 (2020) 96–102

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Microbiome changes and oxidative capability of an anaerobic PCB dechlorinating enrichment culture after oxygen exposure Bruna Matturroa, Giuseppe Mascolob, Simona Rossettia,* a b

Water Research Institute, IRSA-CNR, Via Salaria km, 29,300, Monterotondo, (RM), Italy Water Research Institute, IRSA-CNR, Via F. De Blasio, 5 Bari, (BA), Italy

A R T I C LE I N FO

A B S T R A C T

Keywords: Polychlorinated biphenyls In situ bioremediation Microbiome Marine sediment

Marine sediments may represent a sink of persistent organic pollutants including polychlorinated biphenyls (PCBs), toxic compounds prone to reductive or oxidative biodegradation pathways depending on the degree of chlorination and the positions of the chlorine atoms on the biphenyl rings. Superficial marine sediments can be subjected to episodic sediment resuspension by boat traffic and wind action causing the exposure of the underlying anaerobic layer to oxygen. Under these dynamic conditions, a deeper knowledge of the adaptation capability of the autochthonous microbial communities towards severe changes of the reaction environment is required. Insights into the metabolic potential of sediment community members may contribute greatly to the definition of efficient and reliable in situ bioremediation strategies. In this study, an anaerobic PCB-dechlorinating microbial consortium, developed from the chronically polluted marine sediment of Mar Piccolo (Taranto, Italy), was used to evaluate the response of the sediment microbiome to the imposition of aerobic conditions after prolonged anaerobic incubation. Compared to the anaerobic control, a dramatic change in microbiome composition, with a marked increase of Alphaproteobacteria of up to 39.2 % of total operational taxonomic units (OTUs) was revealed by high-throughput 16S rRNA gene sequencing. Accordingly, a decrement of low chlorinated PCBs (up to 58.3 ± 7.5 % for PCB 18) and the concomitant appearance of genes coding for PCB-degrading biphenyl dioxygenase (bph) were observed at the end of the aerobic incubation, suggesting the occurrence of oxidative PCB biodegradation processes.

1. Introduction Polychlorinated biphenyls (PCBs) are anthropogenic chemicals that exist as multiple isomers (congeners), characterized by different degrees of chlorination and chlorine position. As stable molecules, they have been used in hundreds of industrial and commercial applications (electronic devices, batteries, motors and hydraulic systems, electrical equipment, transformers and capacitors). Since 1970, PCBs have been banned owing to their persistence into the environment and toxicity for humans and wildlife [1]. To date, the widespread distribution of PCBs still represents a considerable environmental concern [2]. Similar to other organic halogenated compounds, such as chlorinated aliphatic hydrocarbons (CAHs), biological processes are known to play a major role in the removal of PCBs [3]. Specialized bacteria are able to dechlorinate them through two possible degradative pathways: anaerobic reductive dehalogenation and aerobic degradation,

depending on the degree of chlorination as well as the position of chlorine atoms on the biphenyl rings [4]. Anaerobic reductive dechlorination occurs primarily on the meta- and para- positions of the highly chlorinated congeners allowing their transformation into less chlorinated forms that generally have lower toxicity and bioaccumulation potential and greater susceptibility to subsequent oxidation by aerobic bacteria [5–7]. A range of the resulting low-substituted CBs can be co-metabolized through oxidative pathways by aerobic bacteria [8,9]. To date, anaerobic bacteria capable of deriving energy from anaerobic reductive dehalogenation of haloaliphates or haloaromatics with hydrogen as the sole electron donor are limited to Dehalococcoides mccartyi (phylum Chloroflexi) and, to a lesser extent, Dehalobacter spp. (phylum Firmicutes), the latter being able to dechlorinate tetrachloroethene, trichloroethene and hexachlorocyclohexane [10]. “Dehalococcoides” species represent a deeply branching lineage within the

Abbreviations: Bph, biphenyl; DAPI, 4′6-diamidino-2-phenylindole; DL, dioxin-like congeners; dw, dry weight; GC-MS, Gas chromatography-mass spectrometry; LOD, limit of detection; LOQ, limit of quantification; NGS, Next Generation Sequencing; OTUs, operational taxonomic units; PCB, polychlorinated biphenyls ⁎ Corresponding author. E-mail address: [email protected] (S. Rossetti). https://doi.org/10.1016/j.nbt.2019.12.004

Available online 24 December 2019 1871-6784/ © 2019 Elsevier B.V. All rights reserved.

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phylum Chloroflexi, and include several uncultured PCB-dechlorinating bacteria such as o-17 and DF-1 strains [11–14]. These uncultured dechlorinating Chloroflexi have been designed as “Dehalococcoides”-like group (DLG) mostly found in PCB contaminated marine or estuarine sediments [11,12,14–18]. Aerobic PCB degradation involves the activity of oxygenases via the biphenyl (bph) catabolic pathway, which includes four reactions (BphA, dioxygenase; BphB, dehydrogenase; BphC, dioxygenase; BphD, hydrolase [5]). Bacteria utilizing the bph pathway co-metabolize a variety of PCB congeners that are transformed in a strain dependent mode [19]. Various species of aerobic bacteria have been described in soil to oxidatively degrade PCBs. They mostly belong to the Gammaproteobacteria and Betaproteobacteria (Pseudomonas, Alcaligenes, Burkholderia, Acinetobacter, Comamonas, Corynebacterium, Rhodococcus and Bacillus) [20]. The majority are involved in the first step of aerobic biphenyl ring cleavage, generating a cis-dihydrodiol intermediate and chlorobenzoate, the latter used by chlorobenzoate-degrading bacterial consortia required for the complete PCB mineralization [20–23]. This capability may be acquired through bph horizontal gene transfer, being these genes located on plasmids [24]. Although bioremediation is an attractive strategy for clean-up of PCB contaminated soil and sediments, it may be limited by low PCB availability, low expression of catabolic genes and consequently difficulties in ring cleavage of the biphenyl molecules [25]. Moreover, because natural biodegradation of PCBs generally occurs at low rates, various solutions to accelerate it have been evaluated [26,27]. Among others, the anaerobic reductive dechlorination of highly chlorinated congeners followed by the aerobic degradation of the products has been proposed as an effective treatment strategy for PCB contaminated sediments [3]. However, studies based on sequential anaerobic-aerobic conditions are limited [28] and mostly conducted in closed soil microcosms with Aroclors® mixtures comprising well-known percentages of tri- and tetra- ortho CBs that are recalcitrant to aerobic degradation and infrequently found in the environment [29]. PCBs in the environment are presented as complex, and often incompletely defined, mixtures. While significant advances have been made in recent years to elucidate the biochemical basis of aerobic PCB biodegradation, difficulties still remain, linked to the analysis of biodegradation performance in PCB congener mixtures as well as the complex metabolic networks responsible for PCB degradation. Moreover, as marine sediments characterized by expansive microbial diversity and considered to be hot spots of element cycling including sulfur and nitrogen, the study of such environmental matrices in terms of microbiome composition under stressed conditions (contamination, change of redox conditions) is of interest [3]. Hence, research efforts describing the microbial communities able to reduce PCB contamination under different reaction conditions are required to define efficient and reliable in situ bioremediation strategies of contaminated matrices, such as superficial marine sediments, where variable redox conditions may exist across the depth or be caused by sediment resuspension. Here, microbiome changes are explored after the introduction of aerobic conditions onto an anaerobic PCB dechlorinating consortium, developed from a PCB chronically contaminated marine sediment and maintained active and stable under reducing redox conditions.

Table 1 Main PCB congeners analyzed in the inoculum and at the end of aerobic and anaerobic incubation. The percent decrement of low chlorinated PCB congeners observed under oxidative conditions was estimated versus the anaerobic control values. PCB ng g−1 dry sediment

Inoculum

Experimental Aerobic

Anaerobic control

Decrement (%)

PCB PCB PCB PCB PCB

32.3 ± 1.2 89.2 ± 9.1 230.7 ± 60 505 ± 140 6.9 ± 0.3

5.7 ± 0.1 30 ± 1.7 29.9 ± 0.8 72.3 ± 8.6 2.3 ± 0.2

13.9 ± 2.7 44.2 ± 3.2 49.3 ± 12 110.2 ± 33 3.6 ± 0.2

58.3 31.7 37.7 32.5 35.8

18 28 + 31 44 52 77

± ± ± ± ±

7.5 8.9 13.5 12.4 2.7

Aliquots of the anaerobic PCB-dechlorinating microbial enrichment (here referred to as the inoculum) were used to prepare duplicate aerobic microcosms in 120 mL serum bottles, with 30 mL of inoculum and 40 mL of synthetic marine water, consisting of NaCl 22 g L− 1, MgCl2·6H2O 9.7 g L− 1, Na2SO4 3.7 g L− 1, CaCl2 1 g L− 1, KCl 0.65 g L− 1 , NaHCO3 0.2 g L− 1, H3BO3 0.023 g L− 1. Aerobic microcosms were prepared without the addition of PCBs, as the contaminants were already present in the inoculum and derived from the previous reductive dechlorination activity (Table 1). Aerobic bottles were maintained under rotation at room temperature for 300 d. Duplicate microcosms were maintained anaerobically (hereinafter referred to as “anaerobic control”), by sealing with Teflon-faced butyl rubber stoppers and flushing with a mixture of N2/CO2 without the addition of the electron donor. 2.2. PCB quantification The extraction and clean up was performed using pressurized solvent extraction (PLE) by an ASE-300 instrument (Thermo Fisher Scientific, Waltham, MA, USA) and 66 mL extraction cells. One half of the cell (lower part) was filled with silica gel previously mixed with H2SO4. The upper half part of the cell was filled with the 5 g freezedried sample mixed with diatomaceous earth and copper in the ratio 1:3:3 which was also spiked with the correct amount of labelled internal standard (13C-PCB104). The extraction was carried out with nhexane (3 cycles at 140 °C). The extract was evaporated to 0.5 mL and a 1 μL sample analyzed by GC/MS-MS (Agilent 7890 GC, 7000C MS and GERSTEL MPS2 autosampler) equipped with an electron impact source (70 eV) and a Agilent DB5-MS 60 m x0.25 mm, 0.25 μm column. The following conditions were used: injector temperature: 270 °C, flow (He): 1.4 mL min−1, initial column temperature: 80 °C (held for 1 min), first ramp: 25 °C min−1 up to 185 °C (held for 1 min), second ramp: 2 °C min−1 up to 210 °C (held for 10 min), third ramp: 5 °C min−1 up to 310 °C (held for 5.3 min). The following PCB congeners were detected and quantified according to their mass transitions (the dioxin-like congeners (DL) are also indicated): PCB18, PCB28 + 31 (these congeners have identical retention times), PCB44, PCB52, and PCB77 (DL). The mass-transitions used for the quantitation method were as follows: 256-186 (PCB18), 256-186 (PCB28 + 31), 255-120 (PCB44), 290-220 (PCB52 and PCB77), 336-266 (13C-PCB104). Quantification was based on a 5-point calibration curve in which each solution contained the congeners as well as the labelled internal standard, and data were reported as ng PCBs per g dry sample. The analytical method was validated with a set of 5 samples as the initial precision and recovery procedure. The limit of detection (LOD) and limit of quantification (LOQ) were evaluated employing native standard spiked samples that were used for determining initial precision and recovery. LOD was evaluated as the concentration reaching a signal 3-fold higher than noise. LOQ was evaluated as 3-times LOD. The LOD of the procedure was determined to be 0.002 ng g−1 for each congener. Quality Control (QC) steps were taken in order to monitor and ensure precision and accuracy. QC practice included analysis of QC

2. Material and methods 2.1. Microcosm set up The PCB degrading enrichment cultures were previously constructed with the contaminated marine sediment from Mar Piccolo (Taranto, Italy - sampling station 11 40° 28′ 46 N, 17° 15′ 38 E) heavily contaminated mostly by penta-CBs, hexa-CBs, and hepta-CBs (∼4.2 mg g−1) [15]. The PCB contaminated sediment was maintained anaerobically under conditions promoting reductive dechlorination over a long-term period [15,16]. 97

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2.7. 16S rRNA Gene Clone Library and phylogenetic analysis

samples with each set of experimental samples, i.e. calibration standards, certified reference materials (where available), spiked samples, duplicate sample analysis and blanks. Additional QC measures included the analysis of blind duplicate samples. Quality Assurance (QA) activities included the review of analytical methods, evaluation of nonconformances as well as review of quality data.

DNA was amplified using primers 27 F (5′-AGAGTTTGATCMTGGC TCAG-3′) and 1492R (5′-TACGGYTACCTTGTTACGACTT-3′) using Hot Start Taq98 (Lucigen, Italy). PCR reactions were performed as follows: 2 min at 98 °C, 38 cycles for 30 s at 98 °C, 30 s at 58 °C, 1 min at 72 °C and final 15 min at 72 °C. PCR products were purified using the QIAquick® PCR purification kit (Qiagen, Milan, Italy) and cloning was carried out using pGEM-T Easy Vector System (Promega, Italy) into E. coli JM109 competent cells (Promega, Madison, WI, USA) according to the manufacturer’s instructions. Positive inserts were amplified from recombinant plasmids obtained from white colonies by PCR using the sequencing primers T7F (5′-TAATACGACTCACTATAGGG-3′) and M13R (5′-TCACACAGGAAACAGCTATGAC-3′). PCR amplicons of 1465 bp length were purified using the QIAquick PCR purification kit and sequenced with primers 530 F (5′-GTGCCAGCMGCCGCCG-3′-) and 907R (5′−CCGTCAATTCMTTTRAGTTT-3′). A total of 100 and 97 clones respectively were screened by PCR for the aerobic and the anaerobic microcosms. Multiple sequence alignments were performed with ClustalW2 to check sequence similarities. The 16S rRNA sequences were analyzed with the ARB software [34] using the SILVA 16S rRNA SSU RefNR99 (release 132, December 2017). Sequences were screened for chimeras using DECIPHER [35] and aligned with ClustalW. The phylogenetic tree was constructed using the maximum likelihood method implemented in the program RAxML [36]. Bootstrap analyses were conducted using 1,000 resampling replicates. Chloroflexi and Elusimicrobia were chosen as outgroups for the phylogenetic trees constructed with 16S rRNA gene sequences from aerobic and anaerobic microcosms, respectively. GenBank accession numbers MK478441, MK478442, MK478443, MK478444, MK478445, MK478446, MK478447, and MK478448 for 16S rRNA gene sequences obtained from the anaerobic control and MK141022, MK141023, MK141024, MK141026, MK141027, and MK141028 for those from the aerobic microcosms.

2.3. Microbial community quantification The analysis was conducted on cells extracted from sediment samples collected from each microcosm before and after treatment. The sediment slurry (1 g dry weight (dw) was collected from the serum bottles with sterile spatulas and immediately fixed in formaldehyde (2 % vol/vol final concentration). Samples were processed to extract cells as previously described [15]. Total cells were stained in Vectashield Mounting Medium® with DAPI (4′,6-diamidino-2-phenylindole, Vector Laboratories, Italy). Cell counting was performed by microscopic analysis on at least 20 randomly selected fields for each sample. Cell abundance was expressed as cells per dw of marine sediment (cells g−1). Means and standard deviations were calculated with Microsoft Excel®. 2.4. DNA extraction DNA was extracted from 0.5 g of dry sediment taken from the anaerobic PCB-dechlorinating enrichment culture utilized for microcosm construction and from either the aerobic experimental microcosms or the anaerobic control after 300 d incubation. DNA was extracted with DNeasy PowerSoil Kit (Qiagen, Hilden, Germany) following manufacturer’s instructions and stored at −20 °C prior to further analysis including PCR, NGS and clonal analysis. 2.5. Polymerase chain reaction (PCR) PCR was performed with the bphA 463 F/674R primer set targeting the functional gene bphA coding for a subunit of PCB-degrading biphenyl dioxygenase [30]. Reactions were performed in a 25 μL total volume including 12.5 μL of Hot Start Taq98 (Lucigen, Middleton, WI, USA), 0.5 μM of each primer and 1 μL DNA, as follows: 2 min at 98 °C, 38 cycles for 30 s at 98 °C, 30 s at 55 °C, 1 min at 72 °C and a final cycle for15 min at 72 °C.

3. Results and discussion 3.1. Evaluation of PCB decrement and bacterial growth A PCB contaminated marine sediment, previously maintained under anaerobic conditions and enriched for PCB dechlorinating bacteria [15, 16] was exposed to aerobic conditions. The residual PCB contamination after the previous reductive dechlorination activity was 864.1 ng PCB g−1 dry sediment (“Inoculum” in Table 1). After the shift to aerobic conditions, a decrement of low chlorinated congeners was observed in the exposed microcosm. At the end of the aerobic treatment, the decrease in tri-CBs and tetra-CBs ranged between 31.7 % and 58.3 % of the concentration measured in the anaerobic control (Table 1). A 5-fold increase of total DAPI stained cells up to 5.61 × 108 ± 8.97 × 107 cells g-1 dry sediment was found at the end of the aerobic incubation compared to the cell abundance estimated at the beginning of the treatment (1.06 × 108 ± 1.37 × 107 cells g-1 dry sediment). In contrast, no significant increase in microbial growth was observed in the anaerobic control (1.21 × 108 ± 1.87 × 107cells g-1 dry sediment; p > 0.05).

2.6. Next generation sequencing (NGS) 10 ng DNA extracted from each sample was used for NGS analysis. 16S rRNA Amplicon Library (V1–3) was prepared as detailed in [14]. PCRs were performed in duplicate in 25 μL reaction volume containing Phusion Master Mix High Fidelity (Thermo Fisher Scientific, Waltham, MA, USA) and 0.5 μM final concentration of the library adaptors with V1–3 primers (27 F: 5′-AGAGTTTGATCCTGGCTCAG-3′; 534R: 5′-ATT ACCGCGGCTGCTGG-3′). The libraries were purified using the Agencourt® AMpure XP bead protocol (Beckmann Coulter, Italy) and concentrations were measured with Qubit 3.0 fluorometer (Thermo Fisher Scientific). The purified libraries were pooled in equimolar concentrations and diluted to 4 nM. 10 % Phix control library was spiked in to overcome low complexity issue often observed with amplicon samples. The samples were paired end sequenced (2 × 301 bp) on a MiSeq (Illumina, San Diego, CA, USA) using a MiSeq Reagent kit v3, 600 cycles (Illumina) following the standard guidelines for preparing and loading samples on the MiSeq. NGS secondary data were processed and analyzed using QIIME2 software tools 2018.2 release [31]. The reads were demultiplexed using demux plugin1, denoized, dereplicated and chimera-filtered using the DADA2 algorithm [32]. Taxonomic analysis was based on a Naïve–Bayes classifier trained on 16S rRNA gene OTUs clustered at 99 % similarities within the Silva 132 database [33].

3.2. Microbiome composition of the inoculum and the anaerobic control Epsilobacteraeota (37.6 % of total OTUs) dominated the microbial community of the anaerobic inoculum with OTUs mostly related to the genus Sulfurovum (36 %, Fig. 1). The occurrence of Sulfurovum species has been already documented in marine sediments contaminated by chlorinated compounds [37] and in samples taken from the Mar Piccolo of Taranto analyzed in previous studies [15,16]. The second most abundant OTU detected in the anaerobic inoculum was affiliated to Deltaproteobacteria (17.5 % of total) and in particular to Syntrophobacteraceae (3.54 %), Desulfarculaceae (1.6 %) and 98

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studies indicating these phyla as key players in PCB reductive dechlorination in marine sediments [16]. 3.3. Changes of the marine sediment microbiome after the imposition of aerobic conditions The analysis of the PCB contaminated marine sediment treated under aerobic conditions revealed a remarkably different microbiome composition compared to the anaerobic inoculum. Aerobic incubation strongly promoted the increase of bacterial OTUs affiliated to Alphaproteobacteria (39.2 %) and Gammaproteobacteria (20 %). Proteobacteria are often found abundantly in PCB-contaminated soil and sediments and their increment has been observed in several aerobic microcosm studies [20,30]. Alphaproteobacteria mostly comprised OTUs related to Rhodobacteraceae (14 %), Sphingomonadaceae (4 %), Rhodospirillaceae (3 %) and Sneathiellaceae (2.3 %) families and to lesser extent to other uncultured Alphaproteobacteria (13 %) (Supplementary Table S1). Rhodobacteraceae are considered key-players in biogeochemical cycling. A characteristic feature of this family is the presence of a large number of phototrophic species containing bacteriochlorophyll and phylogenetically mixed with chemotrophs. They conduct many fundamental metabolic processes such as aerobic respiration, anaerobic fermentation, sulfur oxidation, autotrophic carbon fixation, nitrogen fixation, anoxygenic photosynthesis, or hydrogen production in various combinations [44,45]. Among these, OTUs related to members of Roseobacter clade NAC11-7 lineage were previously found in the PCB contaminated marine sediment under aerobic conditions [46]. Roseobacter clade is one such major marine group capable of aerobic anoxygenic phototrophy and of deriving energy from light without producing oxygen. A role for inorganic sulfur oxidation in many coastal and benthic marine environments (e.g., sediments and sulfide-rich habitats), has been recognized in Roseobacter lineages [46]. Phenotypic assays carried out on cultivated microorganisms indicated that many Roseobacter species are capable of using aromatic compounds as primary growth substrates, leading a novel pathway for the aerobic degradation of benzoate in soil [43,46,47]. OTUs related to the Sphingomonadaceae family include several species able to grow and degrade xenobiotic and recalcitrant polycyclic aromatic compounds of natural or anthropogenic origin, making them of interest to bioremediation applications [48]. Within the Rhodospirillaceae family, OTUs related to Rhodospirillales were also found, the latter indicated as promising candidates for PCB degradation through oxidative pathways in phyto/ rhizoremediation studies [49]. Moreover, Rhodospirillales OTUs found in the aerobic marine sediment also included magnetotactic microaerophilic Magnetospira and Magnetovibrio species, isolated from sulfiderich sediments and able to grow chemolithoautotrophically on thiosulfate and sulfide with oxygen as the terminal electron acceptor [50,51]. Gammaproteobacteria were also in evidence in the aerobic microcosm as the second most abundant group of OTUs (20.44 % of total) including members of Ectothiorhodospirales (4 %) and Steroidobacterales (Woeseiaceae_Woeseia) (3 %). Other OTUs were mostly related to Gammaproteobacteria Incertae Sedis or uncultured bacteria. Ectothiorhodospirales OTUs included members of Ectothiorhodospiraceae and Thioalkalispiraceae (including Thiohalophilus species), known to be marine obligate chemolithoautotroph sulfur oxidizing bacteria using reduced sulfur compounds, including thiocyanate, as electron donors with oxygen or nitrite as electron acceptors [52,53]. Members of the Woeseiaceae family (e.g. Woeseia spp.) have been indicated as abundant core members of microbial communities in marine sediments able to assimilate inorganic carbon [54]. This family is heterogeneous and covers a broad physiological spectrum ranging from facultative sulfur- and hydrogen-based chemolithoautotrophy to obligate chemorganoheterotrophy. This might provide adaptations to various biogeochemical settings and possibly explains their occurrence

Fig. 1. OTU abundances obtained by 16S rRNA gene high-throughput sequencing performed on the inoculum and after aerobic (experimental) and anaerobic (control) incubation. Data are presented as the mean (SD ≤ 10 %) of duplicate assessments.

Desulfobacteraceae (5.3 %) families that include strictly anaerobic and sulfate-reducing bacteria. They are flexible in terms of catabolic reactions and, as sulfate or sulfite reducers, their presence was shown to promote organohalide respiration of PCB dechlorinating bacteria in enrichment cultures [6]. Chloroflexi members represented ∼ 11 % of total OTUs in the inoculum and the main components were affiliated with Dehalococcoidia (5.3 %) and Anaerolineae (5.5 %) classes. Among these, OTUs related to Dehalobium genus were found, the latter already known as marine PCB dechlorinating bacteria [11,13–16,38–41]. OTUs related to Dehalococcoidia SCGCAB-539-J10, known to be widely distributed in the marine subsurface, were also found [42,43]. The anaerobic control retained almost the same microbiome composition of the initial PCB dechlorinating microbial consortium (Fig. 1). Epsilobacteraeota members were 22 % of total OTUs (Sulfurovum 11 %) and Deltaproteobacteria remained unvaried compared to the anaerobic inoculum (11 %). OTUs related to Chloroflexi represented 9 % with Dehalococcoidia members (Fig. 1). This finding is in line with previous 99

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Fig. 2. Maximum-likelihood dendrogram of the bacterial domain based on comparative analyses of 16S rRNA gene data obtained from PCB contaminated marine sediment after aerobic (a) and anaerobic (b) incubation. Full-length 16S rRNA gene sequences of uncultured clones obtained from the aerobic microcosms (MK141022, MK141023, MK141024, MK141026, MK141027, MK141028) and from the anaerobic microcosms (MK478441, MK478442, MK478443, MK478444, MK478445, MK478446, MK478447, MK478448) are shown in panels a) and b) respectively. The bar represents 10 % estimated sequence divergence.

16S rRNA gene sequences obtained from the anaerobic control occupied different positions in the phylogenetic tree to sequences obtained from the aerobic sediment (Fig. 2b). Most sequences were distantly related ( < 98 % of sequence similarity) to Caldithrix species (clone MK478446, 3.1 % of total clones), candidate division KSB1(clone MK478442, 4.2 %), uncultured Gammaproteobacteria (clone MK478444, 20.6 %; MK478445, 18.5 %), uncultured Deltaproteobacteria (clone MK478443, 17.5 %), uncultured Epsilonproteobacteria (clone MK478447, 19.6 %), uncultured Myxococcales (clone MK478448, 3.1 %), Chloroflexi species involved in reductive dechlorination processes such as Dehalogenimonas, Dehalobium and Dehalococcoides species (clone MK478441, 13.4 %) (Fig. 2b, Supplementary Table S1). Overall, both 16S rRNA gene high-throughput sequencing and clonal analysis revealed marked differences in bacterial composition of the marine sediment under oxidative conditions compared to the anaerobic control. This finding evidenced the intriguing capability of the marine sediment to retain oxidative metabolism, including in putative aerobic PCB degrading microorganisms, after prolonged anaerobic incubation. The occurrence of biphenyl dioxygenase (bphA) genes associated with aerobic PCB oxidation was demonstrated by PCR at the end of the aerobic incubation whereas no amplification was observed in the anaerobic control (Supplementary Fig. S1). This finding is in line with a previous study reporting the co-occurrence of bphA and 16S rRNA genes of putative dechlorinating Chloroflexi in a PCB contaminated harbor sediment, indicating the potential for simultaneous aerobic and

in marine sediments worldwide. 3.4. Phylogenetic analysis 16S rRNA gene clone libraries were constructed on the marine sediment taken at the end of aerobic (experimental) and anaerobic (control) incubations to generate full-length 16S rRNA gene sequences. 100 and 97 clones respectively, from the aerobic and anaerobic microcosms, were analyzed. Sequences differing by 2 % were considered a single relatedness group and representatives were fully sequenced on both strands. A total of 14 full-length 16S rRNA gene sequences was obtained (Supplementary Table S2). In line with NGS data, the majority of clones obtained from the aerobic marine sediment fell into known phyla and were closely related ( > 98 % sequence similarity) to uncultured Rhodobacteraceae (clone MK141023, 35 % of total clones), Oceanibaculum species (clone MK141022, 26 %), Woeseia species (clone MK141024, 17 %), uncultured Acidobacteria (clone MK141026, 8 %), uncultured candidate division BRC1 (clone MK141028, 7 %), and uncultured Acidimicrobiales (clone MK141027, 7 %) (Fig. 2a, Supplementary Table S1). Some Oceanibaculum strains have been previously isolated from an aerobic polycyclic aromatic hydrocarbon-degrading consortium, enriched from a marine sediment [55]. It is noteworthy that most sequences found in this study belonged to uncultured bacteria, suggesting that further investigation on their metabolic features and their role in the aerobic PCB degradation in contaminated marine sediments is desirable. 100

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anaerobic PCB degradation [20]. It is well known that, despite differences in genetic organization, the initial enzymatic steps in the aerobic degradation of many aromatic compounds in different bacterial genera are very similar. BphA, B and C homologous enzymes are found in pathways responsible for the degradation of naphthalene, benzoate and benzene [19]. The enzymes in the PCB pathways may transform not only PCBs and their metabolites, but also related compounds such as monocyclic aromatics [56]. As examples, some bph genes have been reported to be capable of efficiently acting both on toluene and monochlorinated biphenyls [57], or of transforming both naphthalene and PCBs [58]. Previous evidence showed that the capacity to transform biphenyls is not restricted to biphenyl-degrading bacteria and that bacterial growth may occur on biphenyls even in the absence of a biphenyl pathway [19]. Indeed, other contaminants may act as co-substrates influencing the composition and activity of biphenyl pathways, both within individual strains and mixed microbial communities, in contaminated matrices characterized by the presence of complex pollutant mixtures. This is probably the case for the marine sediment used in this study, where a pool of different xenobiotics, including PAHs and PCBs, is present, as found previously [59]. In particular, among PAHs analyzed, phenanthrene, fluoranthene, pyrene, benzoanthracene, chrysene, benzofluoranthene, benzopyrene and benzoperylene were found at concentrations from 80−270 ng g−1 of the sediment [60]. The occurrence of bph genes not strictly associated with strains previously described, indicates that the potential for oxidative metabolism exists within the marine sediment microbiome and is worthy of further investigation.

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