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ScienceDirect Microbial dehalogenation of organohalides in marine and estuarine environments Giulio Zanaroli1, Andrea Negroni1, Max M Ha¨ggblom2 and Fabio Fava1 Marine sediments are the ultimate sink and a major entry way into the food chain for many highly halogenated and strongly hydrophobic organic pollutants, such as polychlorinated biphenyls (PCBs), polychlorinated dibenzo-p-dioxins (PCDDs), polybrominated diphenylethers (PBDEs) and 1,1,1-trichloro2,2-bis( p-chlorophenyl)ethane (DDT). Microbial reductive dehalogenation in anaerobic sediments can transform these contaminants into less toxic and more easily biodegradable products. Although little is still known about the diversity of respiratory dehalogenating bacteria and their catabolic genes in marine habitats, the occurrence of dehalogenation under actual site conditions has been reported. This suggests that the activity of dehalogenating microbes may contribute, if properly stimulated, to the in situ bioremediation of marine and estuarine contaminated sediments. Addresses 1 Department of Civil, Chemical, Environmental and Materials Engineering (DICAM), University of Bologna, Via Terracini 28, 40131 Bologna, Italy 2 Department of Biochemistry and Microbiology, School of Environmental and Biological Sciences, Rutgers University, The State University of New Jersey, 76 Lipman Drive, New Brunswick, NJ 08901, USA Corresponding author: Zanaroli, Giulio ([email protected])

Current Opinion in Biotechnology 2015, 33:287–295 This review comes from a themed issue on Environmental biotechnology Edited by Spiros N Agathos and Nico Boon

http://dx.doi.org/10.1016/j.copbio.2015.03.013 0958-1669/# 2015 Elsevier Ltd. All rights reserved.

Introduction A variety of halogenated aliphatic and aromatic compounds have been produced for close to a century and used in several industrial applications, such as solvents, degreasing agents, biocides, pharmaceuticals, plasticizers, hydraulic and heat transfer fluids, flame retardants and intermediates for chemical synthesis; many other organohalogens are by-products of industrial processes. Their extensive production and use has resulted in a widespread environmental contamination. Aquatic sediments are www.sciencedirect.com

among the most relevant long-term reservoirs for hydrophobic, highly halogenated organic pollutants, including polychlorinated biphenyls (PCBs), polychlorinated dibenzo-p-dioxins (PCDDs) and polybrominated diphenyl ethers (PBDEs), as well as a major entry way of these compounds into the food chain. Under anoxic conditions typical of aquatic sediments a few centimetres below the surface, microbial reductive dehalogenation processes can transform these highly halogenated pollutants into less halogenated end-products, that may be more susceptible to subsequent aerobic oxidative biodegradation and are also often less toxic and less prone to bioaccumulation than the parent compounds. These processes therefore represent a promising way for the sustainable remediation of contaminated sediments, whose management is currently mostly based on expensive and high impact dredging operations or, to a lesser extent, via passive sediment capping. Recent reviews have evaluated the main microbial reductive dehalogenation processes observed for PCBs [1–3], PCDDs [4], chlorinated phenols [5], chlorinated benzenes [6] and chlorinated solvents [7,8]. There is also an increasing knowledge of the dehalorespiring microbes (e.g., [9–11]), as well as some key approaches to stimulating the dehalogenation process [1,3–8,12] in soils or freshwater habitats, or by pure or highly enriched cultures in defined media. Much less is known about the occurrence of dehalogenation processes in marine and estuarine sediments [13], which are ultimate sinks for polyhalogenated organic pollutants entering the environment. In addition, the marine environment is a rich source of biogenic organohalogen compounds [14] that may favor the selection of native organohalide respiring bacteria. Moreover, marine systems have distinct geochemical characteristics, among which the high salinity and sulfate concentrations (20–30 mM versus 100–200 mM in freshwater systems) lay a key role in shaping the indigenous microbial communities [15], for example favoring sulfate reduction over methanogenesis as the dominant process for carbon turnover. In estuarine environments the large variation in salinity and sulfate concentration may occur over time and space, leading to highly dynamic and variable redox conditions and microbial activities [16]. Several studies have shown that sulfate impedes the enrichment of dehalogenating communities through interspecific competition for electron donors or through direct inhibition of dehalogenation [13,17]. Understanding how alternate respiratory electron acceptors and the competition for electron donors impact the activity of Current Opinion in Biotechnology 2015, 33:287–295

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dehalogenating bacteria in marine and estuarine sediments is of keen interest. The aim of this paper is to review the major findings available in the recent literature on the microbial dehalogenation of organohalides in marine and estuarine environments, with specific emphasis on first, the potential occurrence and main features of dehalogenation of PCBs, PCDDs, PBDEs, DDT (Figure 1) under actual site or site-mimicking conditions (Table 1), including weathered contaminants, second, the diversity of microbes and their putative catabolic genes detected in marine and estuarine environments, and third, the approaches to stimulate their activity under marine or estuarine in situ-like laboratory conditions.

Detection and characterization of reductive dehalogenation processes in marine systems Several studies demonstrated that anaerobic bacteria of estuarine [18–25] and marine sediments [22,26] from different continents and polluted areas can dechlorinate Figure 1

Cl

Cl

Polychlorinated biphenyls (PCBs) O Cl

Cl

O Polychlorinated dibenzo-p-dioxins (PCDDs) O Br

Br

Polybrominated diphenyl ethers (PBDEs) Cl Cl

Cl

Cl Cl 1,1,1-trichloro-2,2-bis(p-chlorophenyl)ethane (DDT) Current Opinion in Biotechnology

Chemical structures of the main man-made organohalides accumulating in marine sediments. Current Opinion in Biotechnology 2015, 33:287–295

single congeners or commercial mixtures of PCBs in slurry microcosms, as well as in sediment-free enriched cultures [20,27,28], thus indicating that the catabolic capabilities for PCB dechlorination are widespread in estuarine and marine environments. Similar activities were also observed in a few studies undertaken so far on weathered PCBs or carried out under in situ-like biogeochemical conditions (Table 1). Extensive meta and para dechlorination of tetrachlorinated through heptachlorinated weathered PCBs to trichlorinated and dichlorinated congeners has been observed in estuarine sediments of New Bedford Harbor under methanogenic conditions [29]. Slow in situ dechlorination of weathered PCBs was observed in coastal sediments nearby the Sado estuary, where decreasing concentrations of highly chlorinated congeners, coupled to increasing concentrations of lower chlorinated PCBs was detected at increasing sediment depths [30]. This suggests that the higher sulfate concentration in the upper sediment layers may have inhibited the activity of dechlorinating bacteria. Extensive meta and para dechlorination of weathered and spiked PCBs was also documented in marine sediments collected from different locations of the Venice Lagoon suspended in the presence of the site water, either concomitantly with sulfate reduction [31,32] or after complete consumption of sulfate had occurred [33,34]. Faster dechlorination of spiked PCBs was observed in Venice Lagoon slurry cultures enriched in the presence of the (sulfate-rich) site water, although dechlorination rates generally decreased at increasing sulfate reduction rates [35,36]. These studies would thus suggest that microbial reductive dechlorination of weathered PCBs might occur in contaminated marine and estuarine sediments under the in situ biogeochemical conditions. However, high sulfate concentrations might delay or slow down PCB dechlorination, probably as a consequence of competition for electron donors between PCB dechlorinating and sulfate-reducing bacteria. On the other hand, some sulfate reducers are apparently producers of growth factors for PCB dehalorespirers [37], suggesting that complex interactions between these two groups of microbes may affect the onset and rate of PCB dechlorination in estuarine and marine sediments. A rapid and extensive dechlorination of weathered PCBs under actual site conditions was reported for mesocosms of Baltimore Harbor estuarine sediments only upon bioaugmentation with an allochthonous PCB dehalorespiring bacterium [38,39]. Conversely, in estuarine sediments from a tidal marsh in Brunswick (Georgia, USA) incubated with site water amended with electron donors, rapid dechlorination was observed only for a spiked pentachlorobiphenyl congener but not for the weathered PCBs or spiked nonachlorobiphenyls [40], suggesting that, at this site, the low bioavailability of weathered contaminants was the main factor limiting their dechlorination in situ. In general, a number of different PCB dechlorination ‘patterns’, differing in the position of the removed chlorines on the www.sciencedirect.com

Microbial dehalogenation in marine & estuarine habitats Zanaroli et al. 289

Table 1 Reductive dehalogenation processes of weathered organohalides in marine and estuarine sediments and/or under actual site mimicking conditions Site

Sediment feature

New Bedford Harbor (MA, USA) Venice lagoon (Italy)

Marine

Brunswick (GA, USA)

Estuarine

Microcosms, filter sterilized site water

Er-Jen estuary (Taiwan)

Estuarine

Baltimore Harbor (MD, USA)

Estuarine

Sado estuary (Portugal) Passaic estuary (NJ, USA)

Estuarine Estuarine

Palos Verdes (CA, USA) Hailing Bay (PRC) Southern California Bight (CA, USA)

Marine Marine Marine

Microcosms, filtered site water +/ additional sulfate, +/ organic acids Mesocosms, site water, bioaugmented +/ lactate, +/ Fe(0) In situ analyses Microcosms, site water +/ organic acids, +/ H2 In situ analyses In situ analyses In situ analyses

Marine

Experimental conditions

phenyl ring and relative to other chlorines substituents, have been observed. However there is no distinct dechlorination pattern typical of either freshwater or marine/ estuarine environments. Together, these studies indicate that PCB dechlorination may slowly occur in situ in marine and estuarine sediments, but that the low bioavailability of weathered PCBs, the scarcity or low activity of the indigenous PCB dehalorespirers, the competition of these with sulfate-reducers for electron donors, or the combination of these factors, may limit the onset and/or extent of PCB dechlorination. The microbial dehalogenation of PCDDs has been documented in freshwater habitats but only poorly investigated for more saline environments, mostly under welldefined laboratory conditions. Lateral dechlorination of spiked 1,2,3,4-TCDD to 1,2,4-TrCDD, followed by peri dechlorination to 1,3-DCDD, was reported in San Diego Bay estuarine sediments incubated in sulfate-free and sulfate-amended synthetic media [41–43]; the authors speculated that the most toxic congener 2,3,7,8-TCDD would not be formed from higher chlorinated dioxins under such conditions. On the contrary, weathered octa-CDD was shown to undergo peri dechlorination to 2,3,7,8-TCDD, followed by further lateral dechlorination, in contaminated sediments of the Passaic River estuary incubated in the site water amended with organic electron donors [44]. This indicates that at least a transient accumulation of 2,3,7,8-TCDD occurs from sedimentcarried PCDDs under actual site conditions. In addition, www.sciencedirect.com

Organohalide contaminant

Microcosms, marine mineral medium +/ organic acids Microcosms, site water +/ Fe(0) Enrichment slurry culture, site water +/ organic acids and H2

Reference

Weathered PCBs +/ spiked Aroclor 1242 Weathered PCBs +/ spiked 2,3,4,5,6-CB +/ spiked coplanar PCBs Spiked Aroclor 1260 Spiked coplanar PCBs Spiked Aroclor 1260 Weathered PCBs +/ spiked 2,3,4,5,6-CB +/ spiked nona-CBs Spiked coplanar PCBs

[29]

Weathered PCBs

[38] [39] [30] [44]

Weathered PCBs Weathered PCDDs Weathered DDE Weathered DDT and DDE Weathered DDT, DDD, DDE and DDA

[31] [32] [33] [34] [35] [36] [40]

[74]

[50] [51] [52]

in cultures enriched from the same sediment in sulfatefree or sulfate-amended estuarine mineral medium with spiked hepta-CDDs the molar ratio of formed 2,3,7,8TCDD to non-2,3,7,8-TCDDs was significantly lower than ratios observed in freshwater systems, and that lower amounts of 2,3,7,8-TCDD were formed in the presence of sulfate-reduction [45,46]. This led to the conclusion that formation of 2,3,7,8-TCDD from higher chlorinated PCDDs would probably occur to a lesser extent in sediments exhibiting estuarine and marine characteristics. The few reports indicate that different PCDD dechlorination activities may occur depending on the PCDD congeners and the culture conditions employed; more investigations are thus required to evaluate the actual biotransformation of PCDDs in different estuarine and marine contaminated sediments under actual site conditions. The anaerobic microbial dehalogenation of polybrominated diphenyl ethers (PBDEs) has been poorly investigated so far. To the best of the authors’ knowledge, no reports exist on marine or estuarine sediments, except for a recent study that documented the reductive debromination of a spiked hexa-BDE (BDE-153) in a marine sediment, five tidal mangrove marine sediments and two freshwater pond sediments incubated in synthetic media [47]. Reductive debromination was more extensive in mangrove sediments (up to approximately 98% after 90 days), followed by the marine sediment and the freshwater pond sediments. Penta-brominated and Current Opinion in Biotechnology 2015, 33:287–295

290 Environmental biotechnology

tetra-brominated products accumulated in all microcosms, while tri-BDE and di-BDE were also generated in the most active mangrove sediment microcosms. Despite partial debromination of the higher brominated congeners can lead to the accumulation, in synthetic media, of more toxic tetra-BDE and penta-BDE, this study evidenced that marine microbial communities might have the potential to debrominate PBDEs. Sequential reductive dehalogenation of the spiked flame retardant tetrabomobisphenol A (TBBPA) to bisphenol A (BPA) under methanogenic and sulfate-reducing conditions has been demonstrated in anaerobic estuarine sediment cultures prepared with mineral medium [48]. Further studies are required to assess the potential fate of weathered PBDEs and other brominated flame retardants in actual contaminated marine sediments under in situ-like conditions.

profiles of DDE and DDMU across an 11 year time span indicated decreasing DDE and increasing DDMU concentrations at increasing sediment core depths, with differences between top and deeper layers being much more pronounced a decade later [50]. It is possible that the lower sulfate concentration in the deeper sediment layers may have favored the dechlorination activity. The predominant DDT degradation pathway detected in tropical China coastal area, as determined by on-site withdrawals and analyses, involves the following transformations (Figure 2): p,p0 -DDT ! p,p0 -DDD ! p,p0 DDMS (1-chloro-2,2-bis( p-chlorophenyl)ethane) and (2,2-bis( pp,p0 -DDE ! p,p0 -DDMU ! p,p0 -DDNU chlorophenyl)ethylene) [51]. Therefore, microbial dechlorination of DDT and its derivatives in marine sediments may proceed more extensively than observed for non-marine environments. An alternative pathway for DDT degradation has been also proposed in the marine sediments of the southern California Bight [52] according to both sediment analyses and laboratory slurry cultures, proceeding from DDT through DDA (2,2-bis(4-chlorophenyl)acetic acid) to DBP (bis(4-chlorophenyl)methanone), an end-product considered to be more easily biodegraded (Figure 2). This proposed pathway suggests that monitoring the formation of the relatively water soluble DDA might be crucial in determining the fate of DDT and the toxicity and persistence of its derivatives in contaminated sediments under actual site relevant conditions. However, further investigation is required to elucidate the complete DDT degradation pathway and clarify the underlying mechanisms.

The microbial reductive dechlorination of 1,1,1-trichloro2,2-bis( p-chloropenyl)ethane (DDT), 1,1-dichloro-2,2bis( p-chlorophenyl)ethylene (DDE) and 1,1-dichloro2,2-bis( p-chlorophenyl)ethane (DDD) has been reported to occur in marine sediments (Table 1 and Figure 2). The process mostly involves the aliphatic chloroethyl group of the molecule, as observed also in non-marine environments [49], where DDT dechlorination is apparently limited to its transformation to DDD or DDE as end products. Evidence of in situ microbial dechlorination of p,p0 -DDE to p-p0 -DDMU (1-chloro-2,2-bis( p-chlorophenyl)ethylene) was observed in Palos Verde Shelf marine sediments (Figure 2), where vertical concentrations Figure 2 Cl

Cl

Cl

Cl

Cl

Cl

Cl

DDMS

DDD Cl Cl Cl

Cl

Cl

Cl

Cl

DDT

Cl

Cl

Cl

Cl

Cl

DDE

Cl

DDMU HO

O

Cl

O

Cl

DDA

Cl

DDNU

Cl

Cl

DBP Current Opinion in Biotechnology

Dechlorination pathways of DDT detected in marine sediments. Current Opinion in Biotechnology 2015, 33:287–295

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Microbial dehalogenation in marine & estuarine habitats Zanaroli et al. 291

Marine organohalide respiring bacteria and reductive dehalogenase genes Despite a large number of organohalide respiring bacteria have been isolated from aquifers, freshwaters sediments, soils and sludges, only few have been obtained from marine habitats [53–56] and estuarine sediments [37,57] (Table 2). Among them, none is known to be able to dehalorespire PCDDs, PBDEs or DDT. The facultative dehalorespirers Desulfovibrio dechloracetivorans SF3, Desulfomonile liminaris DCB-M, Desulfovibrio sp. TBP-1 and Desulfoluna spongiiphila AA1 can couple the oxidation of several organic electron donors to the reductive dehalogenation of less hydrophobic, hydroxylated or carboxylated monoaromatic organohalides, including ortho-substituted bromophenols commonly produced by marine organisms [58]. The presence of sulfate (an alternate electron acceptor for all these strains) does not inhibit reductive dehalogenation. Two other isolates are obligate dehalorespirers belonging to the class Dehalococcoidia of the Chloroflexi phylum. Dehalococcoides mccartyi MB uses only hydrogen as electron donor and PCE and TCE as electron acceptors, that are reduced to trans-DCE and cis-DCE. Dehalobium chlorocoercia DF-1 couples the oxidation of either formate or hydrogen to the reductive dechlorination of tetrachloroethene and trichloroethene [59], hexa-chlorobenzene and pentachlorobenzene [60], and PCB congeners with double-flanked chlorines [28,37]. Surprisingly, although strain DF-1 can use only organohalides as electron acceptors, the co-presence of 10 mM sulfate (i.e., approximately one third of the concentration typical of marine water), completely inhibits PCB dechlorination [37]. Abrupt shifts in the composition of Dehalococcoidia populations have been reported with depth in marine sediments, indicating that biogeochemical factors, including the availability of electron acceptors such as sulfate, may affect their distribution

[61]. This would suggest that growth and activity of many dechlorinating bacteria might be hampered in marine and estuarine environments and restricted to low-sulfate niches (e.g., increasing depths below the sediment surface) or to estuarine settings. The bacteria that mediate the dechlorination of BBDEs and DDT in microbial communities of marine or estuarine origin have not been identified. Conversely, a number of studies have been carried out to characterize PCB and, to lesser extent, PCDDs-dechlorinating marine and estuarine communities. With the exception of phylotype DEH10, that is related to the Pinellas group of Dehalococcoides spp., all the PCB-dehalorespiring microorganisms identified so far, by combining selective enrichment with molecular monitoring approaches, in estuarine [21,23] and marine sediment cultures [36] cluster with or very close to the non-Dehalococcoides lineage of the class Dehalococcoidia which is represented by D. chlorocoercia DF-1. Among these, strain o-17, which is able to dechlorinate PCBs, chlorobenzenes and PCE, has also been obtained in co-culture with a Desulfovibrio sp. from Baltimore Harbor estuarine sediments [27,62]. In addition, non-Dehalococcoides Dehalococcoidia have also been generally detected in marine and estuarine microcosms and enrichments in association with PCB reductive dechlorination activities [22,34,35,63]. Only one study investigated the PCDDs dechlorinating community in an estuarine sediment [43], suggesting that two non-Dehalococcoides Dehalococcoidia strains related to the DF-1 lineage were possibly responsible for dechlorination. While PCB-dechlorination and PCDD-dechlorination in freshwater sediments has been mainly ascribed to Dehalococcoides spp. [2,9,10,64,65], non-Dehalococcoides Dehalococcoidia appear to be the most probably, yet largely uncharacterized, dehalorespiring bacteria associated with the degradation

Table 2 Dehalorespiring bacteria isolated from marine and estuarine environments Microorganism

Isolation source

Organohalide electron acceptors

Inhibition of dehalogenation by sulfatea,b

Reference

Desulfovibrio dechloracetivorans SF3 Desulfomonile liminaris DCB-M and DCB-F Dehalococcoides mccartyi MB Desulfoluna spongiiphila AA1

Marine sediment

2-Chlorophenol and 2,6-chlorophenol

n.d.

[53]

Marine sediment

Chlorobenzoates

No (5 mM)

[54]

Marine sediment Marine sponge

n.d. No (2.5 mM)

[55] [56]

Desulfovibrio sp. TBP-1

Estuarine sediment

n.d.

[57]

Dehalobium chlorocoercia DF-1

Estuarine sediment

PCE and TCE 2,4,6-Bromophenol and its debromination intermediates, 3-bromophenol bromophenol, 2-iodophenol and 3-iodophenol Bromophenols with only unflaked orto and para bromines Hexa-chlorobenzene and penta-chlorobenzene, PCE, TCE, PCB congeners (double-flanked chlorines)

Yes (10 mM)

[28,37,59,60]

a b

Concentration tested is given in parenthesis. n.d., not determined.

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of these classes of compounds in marine and estuarine sediments. In addition to the detection, enrichment and isolation of marine organohalide respiring bacteria, recent efforts have focused on detection of marine or estuarine reductive dehalogenase homologue genes (rdhA). Putative rdh gene motifs were detected by PCR in a sulfidogenic bromophenol-degrading consortium enriched from estuarine sediment [66] and in TCDD-dechlorinating enrichment cultures [43]. More extensive investigations have been performed on subseafloor sediments, where molecular ecological studies based on 16SrRNA gene sequences revealed a higher abundance of Chloroflexi phylotypes [67]. PCR-based approaches allowed to detect a number of rdhA sequences in sediments collected at different depths (down to 358 m below the seafloor) from several locations in the Pacific Ocean [68,69]. Subseafloor rdhA homologues were preferentially detected in shallow sediments and had similarities ranging from 33% to 64% with previously reported dehalogenase gene sequences. Conversely, new rdhA homologous sequences affiliated with novel clusters were observed with high frequency at five different depth horizons in sediments off the Shimokita Peninsula of Japan by means of shotgun metagenome libraries, thus revealing a much higher diversity of rdhA homologous genes than previously documented [70]. Although sequence similarity and substrate specificity of known reductive dehalogenases have been shown to be poorly correlated [11] and some reductive dehalogenases were shown to dechlorinate both PCE and PCB congeners [71], this intriguing diversity suggests that the marine environment might harbor reductive dehalogenases with new dehalogenating activities and biochemical properties.

Biostimulation and bioaugmentation under marine in situ mimicking conditions Several approaches, mainly aiming at stimulating the growth and activity of indigenous dehalorespiring bacteria, have been investigated to enhance the reductive dehalogenation of organohalides in marine and estuarine sediments. These studies have focused on PCB and PCDD dechlorination, while no attempt has been made so far to stimulate the reductive dehalogenation of PBDEs or DDT in marine or estuarine environments. ‘Halopriming’, that is, the supplementation of a halogenated alternative electron acceptor, has been shown to remarkably stimulate PCB and PCDD reductive dechlorination of spiked contaminants in marine and estuarine sediment cultures established in synthetic media [19,42,46], although it did not affect the reductive dechlorination of weathered PCBs under in situ-like conditions [32,33,40]. Although brominated haloprimers have the advantage of being often completely dehalogenated and further mineralized by the indigenous microbial community, their priming activity is remarkably lower Current Opinion in Biotechnology 2015, 33:287–295

than that of chlorinated primers with chemical structures more analogous to the target pollutants which are, conversely, only partially dehalogenated to chlorinated endproducts [42]. Halopriming does not appear therefore to be a suitable approach for the in situ stimulation of reductive dehalogenation processes. Contradictory results have been reported on the effects of biostimulation via addition of organic electron donors or hydrogen. Organic acids stimulated the reductive dechlorination of weathered octa-CDDs to hepta-CDDs under in situ-mimicking conditions, while supplementation of hydrogen gas in the microcosms headspace enhanced dechlorination up to mono-CDDs [44]. However, the addition of electron donors did not affect the reductive dechlorination of weathered PCBs in high organic content marine sediments [29] or when supplied to bioaugmented estuarine sediment mesocosms [39]. In addition, repeated feeding with high concentrations of electron donors lowered dehalogenation activity and instead stimulated competing sulfate-reducing and methanogenic metabolisms [36]. Conversely, supplementation of slow hydrogen releasing zerovalent iron particles enhanced dechlorination of spiked PCBs and promoted the enrichment of putative dehalorespinig Chloroflexi in marine sediment microcosms [34,72], while it did not significantly affect weathered PCB dechlorination in bioaugmented estuarine mesocosms [39]. Thus, several site-specific factors, such as the geochemical features of the site, the concentration of indigenous dehalorespirers and the composition of the indigenous microbial community, may affect the effectiveness of these approaches, which probably requires a site-tailored optimization in terms of electron donors type, concentration and feeding frequency. The availability of a pure PCB dehalorespiring culture (D. chlorocoercia DF-1) obtained from estuarine sediments has recently allowed investigation into the effectiveness of bioaugmentation as a potential approach for the remediation of PCB-contaminated sediments. Bioaugmentation of Baltimore Harbor sediment mesocosms overlaid with site water with D. chlorocoercia DF-1 promoted the extensive dechlorination of weathered PCBs within 120 days; while the maintenance of DF-1 within the indigenous population was confirmed for the whole period, dechlorination of both doubly flanked and single flanked chlorines indicated an apparent synergistic effect on the indigenous dechlorinating community [38]. More recently, the co-bioaugmentation of Baltimore Harbor sediment mesocosms with D. chlorocoercia DF-1 and Burkholderia xenovorans LB400 proved effective in promoting the concurrent anaerobic dechlorination and aerobic degradation of weathered PCBs [39]. Remarkably, 80% decrease by mass of PCBs was obtained after 120 days with no significant increase in lesser chlorinated congeners; biodegradation of lesser chlorinated products ended however at day 120, probably as a consequence of the lack of a non-chlorinated cometabolic substrate, while reductive dechlorination proceeded www.sciencedirect.com

Microbial dehalogenation in marine & estuarine habitats Zanaroli et al. 293

for one year. These studies support the feasibility of using in situ bioaugmentation to promote remediation of PCB impacted estuarine sediments. Marine sediments might provide a less favorable environment for B. xenovorans LB400 and D. chlororcoercia DF-1; the selection of PCB degraders coping with higher salinity and of PCB dehalorespirers adapted to higher sulfate concentrations might be required to address bioaugmentation of marine sediments. Bioaugmentation studies with a highly enriched PCB dechlorinating marine culture in actual site sediments of Venice lagoon are currently in progress in our laboratory; preliminary evidences are indicating the effectiveness of the bioaugmentation approach also in marine sediments. Possible strategies for the production of large amounts of inoculum and its deployment to the contaminated sediment, which are required to implement in situ the bioaugmentation of organohalide contaminated sediments, have also been recently proposed [3,38].

Conclusions Robust evidence of the occurrence of microbial reductive dehalogenation in situ or under actual site conditions of contaminated marine and estuarine sediments has been provided mainly for PCBs and DDT and its derivatives, while limited information is available for PCDDs and PBDEs, till date. Although a direct comparison among results for different marine or estuarine sediments and/or different organohalides is difficult to make, the lag phase, the rate and the pathway of dehalogenation appear largely influenced by a combination of factors, such as the type of intrinsic dehalorespiring capability of the indigenous microbes, the bioavailability of the pollutant, and the geochemical features of the sediment. In particular, high concentrations of sulfate arguably delay or slow down the dechlorination process, probably as a consequence of competition for electron donors between dechlorinators and sulfate-reducers; this might be the reason for the limited occurrence and extent of dechlorination processes observed in situ in marine and estuarine surface sediments. Further, the coexistence of other pollutants along with the occurrence of lower and variable temperatures, (bio)turbation phenomena, etc., might additionally affect the occurrence and the route of biotransformation processes under in situ conditions [73]. Different positive effects have been reported upon sediment biostimulation and the few bioaugmentation experiments performed so far under actual site-mimicking laboratory conditions demonstrate the potential for enhancing the degradation of PCBs in contaminated estuarine and marine sediments. Such evidence obtained under site-relevant conditions is encouraging. However, additional information under site/in situ conditions is necessary for an assessment of the actual potential/relevance of in situ natural recovery of sediments and for developing effective and robust protocols for intensifying dehalogenation activities under in situ conditions. www.sciencedirect.com

Acknowledgement Co-funding by the European Commission under Grant Agreement no. 266473 (ULIXES project, 7th FP) is acknowledged.

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