Identification of two organohalide-respiring Dehalococcoidia associated to different dechlorination activities in PCB-impacted marine sediments
© The Author(s) 2017
Received: 23 November 2016
Accepted: 14 July 2017
Published: 24 July 2017
Microbial reductive dechlorination of polychlorinated biphenyls (PCBs) plays a major role in detoxifying anoxic contaminated freshwater and marine sediments from PCBs. Known members of the phylum Chloroflexi are typically responsible for this activity in freshwater sediments, whereas less is known about the microorganisms responsible for this activity in marine sediments. PCB-respiring activities were detected in PCB-impacted marine sediments of the Venice Lagoon. The aim of this work was to identify the indigenous organohalide-respiring microorganisms in such environments and assess their dechlorination specificity against spiked Aroclor™ 1254 PCBs under laboratory conditions resembling the in situ biogeochemistry.
High PCB dechlorination activities (from 150 ± 7 to 380 ± 44 μmol of chlorine removed kg−1 week−1) were detected in three out of six sediments sampled from different locations of the lagoon. An uncultured non-Dehalococcoides phylotype of the class Dehalococcoidia closely related to Dehalobium chlorocoercia DF-1, namely phylotype VLD-1, was detected and enriched up to 109 16S rRNA gene copies per gram of sediment where dechlorination activities were higher and 25-4/24-4 and 25-2/24-2/4-4 chlorobiphenyls (CB) accumulated as the main tri-/dichlorinated products. Conversely, a different phylotype closely related to the SF1/m-1 clade, namely VLD-2, also enriched highly where lower dechlorination activity and the accumulation of 25-3 CB as main tri-chlorinated product occurred, albeit in the simultaneous presence of VLD-1. Both phylotypes showed growth yields higher or comparable to known organohalide respirers and neither phylotypes enriched in sediment cultures not exhibiting dechlorination.
These findings confirm the presence of different PCB-respiring microorganisms in the indigenous microbial communities of Venice Lagoon sediments and relate two non-Dehalococcoides phylotypes of the class Dehalococcoidia to different PCB dechlorination rates and specificities.
Polychlorinated biphenyls (PCBs) are a family of 209 congeners composed by a biphenyl ring carrying one to ten chlorine substitutions. After more than 30 years of worldwide ban, PCBs are still widespread environmental contaminants  and are included in the list of persistent organic pollutants (POPs) targeted for elimination by the Stockholm Convention . The most relevant long-term reservoirs of PCBs released into the environment are aquatic sediments, where PCBs can exert broad toxicity to wildlife  and enter the food chain [1, 4, 5]. Under anoxic conditions, generally occurring in such sediments few centimetres below the surface, highly chlorinated PCB congeners can be transformed by microbial activities, which sequentially remove chlorine atoms from the pollutants, making them be more susceptible to aerobic oxidative biodegradation and often less toxic and less prone to bioaccumulate than parent compounds [6, 7]. Such activity is called microbial reductive dechlorination and represents a promising process for the sustainable remediation of contaminated sediments, currently managed through expensive and highly-impacting dredging operations or capping [8, 9].
The occurrence of microbial PCB-dechlorinating activity in freshwater and estuarine sediments was documented in several laboratory sediment cultures and sediment-free cultures developed with synthetic media [7, 10–16]. The process is mainly mediated by members of the phylum Chloroflexi and, less frequently, Firmicutes, which use PCBs as terminal acceptor for their electron transport chain [7, 10, 17, 18]. Among the organohalide-respiring Chloroflexi, members of the genus Dehalococcoides were predominantly associated to PCB dechlorination in freshwater systems [10, 12, 14, 16, 19, 20] and more rarely in estuarine environments [21, 22]. On the other hand, Dehalococcoides-like Dehalococcoidia belonging to the o-17/DF-1 clade have been more frequently linked to PCB dechlorination activity in estuarine environments [21, 23–25], where salinity and sulfate concentrations may shift over time and space between freshwater and marine conditions . Moreover, a number of pure cultures of Dehalococcoides species isolated from freshwater environments were recently shown to dechlorinate a wider range of organohalides, including PCBs [19, 27–29], in which some reductive dehalogenases were characterized in terms of regio specificities towards PCB congeners [27, 30]. Notably, a non-Dehalococcoides strain, Dehalobium chlorocoercia DF-1, is the sole Chloroflexi PCB respiring isolate obtained so far from estuarine sediments [24, 31].
Microbial dechlorination has been less investigated in marine conditions [17, 32–38] where higher salinity and sulfate concentrations select for different microbial taxa compared to estuarine or freshwater environments . Limited information is also available on the PCB respiring Chloroflexi in marine environments. In sediments of the Gulf of Taranto (Mar Piccolo, Ionian Sea, Italy), the enrichment of Dehalococcoides mccartyi and its PCB reductive dehalogenase-coding genes was reported during PCB dechlorination [40, 41]. Conversely, non-Dehalococcoides Dehalococcoidia were detected in sediments of the Venice Lagoon after sub-culturing of the indigenous community in the presence of exogenous PCBs [34, 42] as well as after biostimulation with zerovalent iron nanoparticles , suggesting that this group of organohalide respiring Chloroflexi might be relevant for the bioremediation of PCB-contaminated marine sediments. The aims of this work were: (i) to further assess the dechlorination potential, in terms of occurrence, dechlorination rate and specificity, of the indigenous microbial communities of the PCB-contaminated sediments in the Porto Marghera area (Venice Lagoon) and (ii) to assess the role of Dehalococcoidia in the process in terms of dehalogenation specificity and growth during organohalide respiration.
Results and discussion
Conversely, 24-25 CB (760 µmol kg−1) and 25-25 CB (511 µmol kg−1) were the most accumulated dechlorination products in sediment C, where formation of 25-3 CB (227 µmol kg−1) and 24-3 CB (105 µmol kg−1) was also detected, along with a much lower accumulation of 246-24/245-3 CB (117 vs 405 µmol kg−1), 25-2/24-2/4-4 CB (100 vs 213 µmol kg−1), 24-24/245-2/246-4 CB (84 vs 510 µmol kg−1) and 25-4/24-4 CB (42 vs 356 µmol kg−1) compared to sediment D (Fig. 2). Considering all the possible dechlorination pathways leading from the mostly depleted congeners of the Aroclor 1254 mixture (i.e., 245-25, 245-34 and 234-245/2346-34 CB) to the dechlorination products that accumulated in higher amounts (Fig. 3b), the most recurrent dechlorination activity in sediment C cultures was directed towards the para position, in particular of single flanked chlorines in 34, 234 and 245 chlorophenyl rings. In addition, some meta dechlorination, mainly of single flanked chlorines in 245 chlorophenyl rings and, to less extent, of double flanked chlorines in 234 chlorophenyl rings, may have occurred in some dechlorination pathway (Fig. 3b). Overall, while para dechlorination of single flanked chlorines in 245 and 34 chlorophenyl rings and meta dechlorination of single flanked chlorines in 245 chlorophenyls are common to potential pathways identified in both sediment C and sediment D cultures, only potential pathways identified for sediment D culture involve also recurrent dechlorination of meta double-flanked chlorines on 2345, 2346 and 234 chlorophenyl rings and of meta and para unflanked chlorines on 24 and 25 chlorophenyl rings. These differences between sediment D and sediment C cultures suggest that PCB dehalogenating bacteria having different dechlorination specificities were enriched in the 2 cultures.
Sediment E1 culture exhibited a dechlorination pattern similar to that of sediment D, being 24-25 CB (879 µmol kg−1), 25-25 CB (514 µmol kg−1), 236-4 CB (348 µmol kg−1), 25-2/24-2/4-4 (230 µmol kg−1) and 24-24/245-2/246-4 CB (225 µmol kg−1) the main dechlorination products accumulated at the end of incubation (Fig. 2). Few common features to sediment C culture were however also observed, such as the detection of 24-3 CB (205 µmol kg-1) and 25-3 CB (88 µmol kg−1) (Fig. 2). A dechlorination pattern apparently similar to that of sediment D culture was observed also in sediment E2 culture, although the less extensive dechlorination observed limits the possibility to compare it with the other sediment cultures (Fig. 2).
Sulfate-reducing and methanogenic activities
Since sulfate-reducing and methanogenic bacteria are the main competitors for electron donors of organohalide respirers in anaerobic sediments, sulfate-reduction and methanogenesis were monitored over incubation to assess the presence of possible inhibitory effects of these metabolisms on the occurrence of PCB respiration. Sulfate concentrations spanned from 27 ± 1 to 39 ± 1 mM at the beginning of the incubation. Sulfate consumption was first detected after a lag phase of 6–11 weeks of incubation except in sediment C cultures, where it started immediately and consumed more than 40% of sulfate in the first 6 weeks (Additional file 2: Figure S1). Complete sulfate depletion was achieved in sediment D and sediment C cultures after 11 weeks of incubation, i.e. before the onset of the dechlorination process, but also in sediment A and F cultures, where no PCB dechlorination occurred (Additional file 2: Figure S1). Slower sulfate reduction took place in the non-PCB-dechlorinating culture B, where sulfate was completely reduced before week 23, and in PCB-dechlorinating sediment E1 and E2 cultures, where complete sulfate depletion occurred at weeks 14 and 23 (Additional file 2: Figure S1). No correlation was therefore observed between the extent and rate of sulfate reduction and the occurrence, extent and rate of PCB dechlorination.
Negligible methane production (lower than 0.2 mmol over 31 weeks of incubation) was observed in all sediment cultures, except for the non-PCB-dechlorinating cultures of sediment A, where 56 ± 4 mL (i.e. 2.5 ± 0.2 mmol) of methane were produced (Additional file 3: Figure S2). Therefore, methane production was not related to the dechlorination activities detected, as previously observed in the same area [35, 42].
Changes in bacterial communities and identification of PCB dechlorinating bacteria
16S rRNA gene copy numbers and relative abundance of phylotypes VLD-1 and VLD-2
Organic chlorine (µmol/gdw)
(16S rRNA gene copies per gram of sediment)
(16S rRNA gene copies per gram of sediment)
Relative abundance (%)
23.2 ± 0.4
0.1 ± 0.0
0.4 ± 0.3
20.4 ± 1.5
2.5 ± 1.9
55.9 ± 5.2
23.2 ± 0.7
0.6 ± 0.6
0.8 ± 0.1
18.6 ± 0.3
15.2 ± 1.0
2.1 ± 0.9
23.3 ± 0.1
1.2 ± 0.4
2.4 ± 1.3
19.4 ± 0.2
55.5 ± 2.6
0.0 ± 0.0
23.3 ± 0.9
0.3 ± 0.1
0.5 ± 0.1
21.0 ± 0.9
58.2 ± 4.8
22.0 ± 1.7
Three out of six marine sediments used in the study showed PCB-dechlorination activities under laboratory conditions resembling the in situ biogeochemistry, suggesting that a potential for dehalogenation is present, although not ubiquitously, in the microbial communities of the Porto Marghera area of Venice Lagoon. Two non-Dehalococcoides phylotypes of Dehalcoccoidia, closely related to other PCB-respiring microorganisms previously identified in estuarine sediments, were associated to two distinct dechlorination activities. Among these, phylotype VLD-1 is capable of unflanked meta and para chlorines removal, and thus potentially able to achieve extensive decontamination and detoxification of sediments impacted by complex PCB mixtures. These non-Dehalococcoides Dehalococcoidia are therefore candidate targets for further enrichment and isolation efforts, aiming at the production of PCB dechlorinating inocula for bioaugmentation purposes, and/or for biostimulation approaches to promote the decontamination of sediments in the Venice Lagoon and other PCB impacted marine areas.
Venice Lagoon sediments
Six sediments (A, B, C, D, E, and F) from different locations of the first industrial area of Porto Marghera (Venice Lagoon, Italy) were used in this study, along with the seawater collected from the same area. Sediments were impacted by PCBs in the range 0.2 (sediment F)–3.3 (sediment D) mg kg−1 consisting predominantly of highly chlorinated congeners (i.e. sediments B and D) but also with considerable percentages of medium–low chlorinated ones (i.e., sediment A) (Additional file 1: Table S1).
Preparation and sampling of sediment cultures
A set of four 100 mL anaerobic slurry cultures (duplicate biologically active and sterile controls) was prepared for each sediment anaerobically using site water for sediment re-suspension (20% dry w/v) under nitrogen atmosphere . Given the limited number and low concentration of PCB congeners occurring in each sediment (Additional file 1: Table S1), sediment cultures were spiked with Aroclor 1254 (20 g L−1 stock solution in acetone) at a final concentration of 1 g of PCBs kg−1, to favour the enrichment of sediment indigenous PCB-respiring bacteria and better assess their PCB dechlorination potential. Sediment cultures were incubated statically in the dark at 28 °C for 31 weeks and periodically sampled according to the procedure described by  to analyse: (i) the volume and composition of the biogas (i.e., methane and CO2), (ii) the congeners and concentration of PCBs in the sediment, (iii) the concentration of sulfate in the water phase and (iv) the structure and composition of the microbial community.
PCB extraction and analytical procedures
PCBs were batch extracted in duplicate from each replicate culture according to procedures described elsewhere . GC-ECD analyses of extracted PCBs were performed under the analytical conditions described in literature . Qualitative analysis was performed by comparing retention times (relative to octachloronaphtalene) of parent PCBs and their dechlorination products with those of PCBs occurring in Aroclor 1242 and Aroclor 1254 standard mixtures (Ultra Scientific Italia, Bologna). Quantitative analyses were performed using the GC-ECD response factor of each PCB, obtained from linear five-points calibration curves Aroclors (in the range 1.0–50.0 mg L−1) and the weight percentage of each congener occurring in Aroclors reported elsewhere . PCB concentrations (μmoles kg−1, referred to the sediment dry weight), the average number of Cl per biphenyl and dechlorination rates (μmoles of Cl released kg−1 week−1, referred to the sediment dry weight) were calculated assuming co-eluting congeners to be present in equal proportions as described in previous works .
Biogas production was measured at each sampling with an airtight glass syringe, while its composition was determined via µGC-TCD as described previously . Sulfate concentration in the water phase was determined with IC-CD as described in .
Community analysis by PCR-DGGE of the 16S rRNA gene
Total DNA was extracted from the wet sediment (approximately 250 mg) recovered from the centrifugation of 2 mL slurry samples at 10,000×g for 10 min with the UltraClean Soil DNA kit (MoBio Laboratories, Carlsbad, CA, USA) according to the protocol “for maximum yields” provided by the manufacturer preceded by treatment with Proteinase K and Lysozyme as described elsewhere . Total DNA was quantified using Qubit® dsDNA HS Assay Kit with a Qubit 3.0 fluorimeter, following the manufacturer’s specifications.
For DGGE analysis, 16S rRNA genes of the bacterial community were PCR-amplified with the GC-clamped primer GC-357f and primer 907r , while 16S rRNA genes of putative dechlorinating Chloroflexi were PCR amplified with the GC-clamped primer GC-348f and primer 844r . PCR reactions were performed as described in previous works . DGGE of PCR products (approximately 400 ng DNA per lane) were performed in 7% (w/v) polyacrylamide gels at 55 V for 16 h. Denaturing gradients from 40 to 60% denaturant and from 45 to 55% denaturant were used to resolve total Bacteria and Chloroflexi-specific amplicons, respectively. Digital images of gels were captured in UV transillumination after staining with SYBR Green I.
Community richness (Rr) and community organization (Co) indexes were calculated from DGGE image analysis as described in literature [50–52]. In particular, the range-weighted richness was calculated from the total number of bands in the pattern and the denaturing gradient comprised between the first and the last band of the pattern, whereas the community organization was derived from Pareto-Lorenz (PL) evenness curves and the respective Gini coefficient.
Most prominent DGGE gel bands were cut, DNA eluted overnight at 4 °C in sterile water, re-amplified and resolved again in DGGE as described above, before amplification with non-GC-clamped primers. Amplicons were finally purified in the presence of 10 U of ExoI and 1 U of FastAP enzymes (Thermo Scientific Italia s.r.l, Milan, Italy) at 37 °C for 15 min before sequencing with the corresponding forward primer. Sequencing was performed by BMR Genomics (Padova, Italy). Each 16S rRNA gene sequence obtained (∼500 bp) was aligned to the bacterial 16S rRNA database of the Ribosomal Database Project (RDP, release 11, http://rdp.cme.msu.edu) and the closest relative and closest cultured relative retrieved with the Seqmatch tool. The phylogenetic affiliation of each sequence was obtained from the same website with the Classifier tool. Nucleotide sequences were deposited in the GenBank database under the Accession Numbers KR013281 and KR013282.
Quantification of target Dehalococcoidia by qPCR
Primer pairs were designed for the specific amplification of two partial 16S rRNA gene sequences detected in the PCB dechlorinating cultures by DGGE using the bacterial primers pair 357f/907r (see Results section). Primers were designed with the Primer-BLAST website (http://www.ncbi.nlm.nih.gov/tools/primer-blast) , using each target sequence as PCR template, a primer melting temperature from 58 to 61 °C, the Refseq RNA database for specificity checking, PCR product Tm between 80 and 90 °C, optimum at 80 °C. The specificity of each candidate primer pairs was checked in silico with the ProbeMatch tool available at the RDP II web site; primer pairs having the highest in silico specificity were checked for cross-specificity in vitro in end-point PCR and qPCR assays (see details below).
The gene copies of the target Chloroflexi phylotypes and the total bacterial 16S rRNA genes were quantified through qPCR using a StepOne™ Real-Time PCR System (Applied Biosystems, Monza, Italy) and StepOne software v 2.0 according to the manufacturer instructions. Designed primer pairs used to target selected Chloroflexi were: (i) 682f (5′-AGGCGAAAGCGGTTTCCAA-3′) and 814r (5′-ACTTAAAGCGTTAGCTTCGGCA-3′) for VLD-1; (ii) 585f (5′- TCAACTGGGAGGAGTCATTCG-3′) and 697r (5′- GAAACAGCCTAGAAAACCGCC-3′) for VLD-2 (Additional file 4: Table S2). Finally, known primers 920f  and 1044r  were used to target bacterial 16S rRNA genes. PCR cycles were as follows: (i) 95 °C for 10 min, (ii) 40 cycles of denaturing at 95 °C for 30 s, annealing at 56, 59 or 54 °C for 30 s (with uncultured Chloroflexi VLD-1, uncultured Chloroflexi VLD-2 and total bacterial 16S rRNA gene primer pairs, respectively), and elongation at 72 °C for 30 secs, (iii) denaturation at 95 °C for 15 s followed by a melting curve from 60 to 95 °C and fluorescence measure every 0.3 °C. The qPCR reactions (25 µL) were set-up as follows: 1× Power SYBR® Green PCR Master Mix (Applied Biosystems, Monza, Italy), forward and reverse primers at 350 nM each and 2.5 µL of DNA template. 5-point standard curves were included in each plate from 6.8 × 103 to 6.8 × 108 copy numbers using PCR products of DGGE gel-purified bands as standard template for Chloroflexi-specific targets or of E. coli 16S rRNA gene for total bacteria. DNA was purified with Wizard® SV Gel and PCR Clean-Up System (Promega Italia S.r.l, Milano, Italy) according to the manufacturer protocol and quantified with a P330 Nanophotometer (Implen GmbH, Munich, Germany). Amplification efficiencies ranged from 85 to 106%, with R2 = 96–100%. Samples and standards were set up in triplicate reactions. Genes copy numbers per gram of sediment were calculated and relative abundance was determined as percentage ratio between the copy number of target Chloroflexi 16S rRNA genes and the copy number of total bacterial 16S rRNA genes obtained from the same template DNA.
ANu performed most of the experiments and carried out the data analyses and interpretation. ANe helped with microcosms set-up and DGGE image-analysis. GZ contributed to the study design and data interpretation. GZ and ANu wrote the manuscript. GZ and FF conceived the study and participated in its coordination. All authors read and approved the final manuscript.
The authors declare that they have no competing interests.
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