Anaerobic microbial dehalogenation and its key players in the contaminated Bitterfeld-Wolfen megasite

Anaerobic microbial dehalogenation and its key players in the contaminated Bitterfeld-Wolfen... Abstract The megasite Bitterfeld-Wolfen is highly contaminated as a result of accidents and because of dumping of wastes from local chemical industries in the last century. A variety of contaminants including chlorinated ethenes and benzenes, hexachlorohexanes and chlorinated dioxins can still be found in the groundwater and (river) sediments. Investigations of the in situ microbial transformation of organohalides have been performed only over the last two decades at this megasite. In this review, we summarise the research on the activity of anaerobic dehalogenating bacteria at the field site in Bitterfeld-Wolfen, focusing on chlorinated ethenes, monochlorobenzene and chlorinated dioxins. Various methods and concepts were applied including ex situ cultivation and isolation, and in situ analysis of hydrochemical parameters, compound-specific stable isotope analysis of contaminants, 13C-tracer studies and molecular markers. Overall, biotransformation of organohalides is ongoing at the field site and Dehalococcoides mccartyi species play an important role in the detoxification process in the Bitterfeld-Wolfen region. in situ, organohalide biotransformation, reductive dechlorination, Dehalococcoides, monochlorobenzene, chlorinated ethenes, chlorinated dibenzo-p-dioxins INTRODUCTION Since the discovery of reductive dechlorination as a respiratory process (Dolfing and Tiedje 1987; Dolfing 1990) and its potential as a remediation strategy for organohalides (Holliger et al. 1997), research has extended from the investigation of the microorganisms and enzymes involved to the identification of these microorganisms and the responsible genes at contaminated field sites. Bioremediation and bioaugmentation strategies to clean up contaminated sites, especially those with chlorinated ethenes, have been implemented and the detection of microorganisms, such as Dehalococcoides, and their reductive dehalogenase genes (rdhA) has become routine (Major et al. 2002; Ritalahti et al. 2005; Stroo, Major and Gossett 2010). Several methods are currently available, usually applied in combination (for an overview see Table 1), for the investigation of the ecophysiology of microorganisms capable of the degradation of organohalides under anoxic conditions. For site assessment, more traditional methods such as analysis of the geochemistry as an indicator of the geochemical conditions, and contaminant and product concentrations as indicators for (bio)transformation are used in combination with newer approaches. These new approaches include compound-specific stable isotope analysis (Hunkeler et al. 2008; Nijenhuis et al. 2016) and detection of biomarkers, i.e. 16S rRNA genes of organohalide-respiring bacteria (OHRB) and their functional genes (Wilson 2010; Nijenhuis and Kuntze 2016). Combined ex situ, laboratory and microcosm approaches can be applied to investigate the potential for biotransformation of a contaminant and elucidation of conditions that improve activity. Stable isotope tracer approaches, applying a 13C-labelled contaminant in in situ microcosms, have been implemented in cases where the contaminant may be used as a carbon source and can be traced into metabolites and microbial biomass (Bombach et al. 2010). This method does not allow direct detection of OHRB as the carbon present in the contaminant is not used as carbon source but an external carbon source, e.g. acetate, is needed (Stelzer et al. 2006; Kittelmann and Friedrich 2008b). Table 1. Indicators/methods applied for the investigation of organohalide degradation in situ (adapted from Bombach et al. 2010). Approach/ concept  Direct/ indirect proof  Indicators and principle  Advantage  Challenges  Examples/references  Hydrochemistry  Redox processes  Indirect  Presence of methane, sulphide and reduced iron (FeII), low redox and absence of oxygen indicate anoxic conditions required for reductive dehalogenation  Easy to analyse (iron, sulphate, nitrate etc.)  Process may be linked to other substrates  Nijenhuis et al. (2007a, 2009, 2013)  Contamination pattern  Direct  Concentration decrease may be a result of degradation but also dilution or sorption  Easy to analyse (GC-FID/MS; HPLC, etc.)  Co-contaminations, decrease of concentration due to sorption, dilution, etc.  Heidrich, Weiß and Kaschl (2004b)  Metabolite detection  Direct  Detection of dehalogenation products such as TCE, DCE, VC and ethene  Proof for pathway  Metabolite may be present as co-contaminant or as degradation product from multiple processes  Heidrich, Weiß and Kaschl (2004b)  Compound-specific stable isotope analysis  Direct  Enrichment of heavy stable isotopes (13C, 2H, 37Cl) indicates degradation of the compound; in reductive dehalogenation correlated with a depleted signature in the final product  Qualitative and potentially quantitative assessment of biodegradation  Equipment not universally accessible and applicable; complementary methods needed for quantitative assessment; reference laboratory studies not always available; isotope compositions of intermediates are difficult to interpret  Kaschl et al. (2005); Imfeld et al. (2008b); Nijenhuis et al. (2016)  Biomarker/molecular biology  PCR/qPCR  Direct  rRNA and rRNA genes of OHRB  Rapid identification of organisms present  Not or only an indirect indicator of activity; only known organisms and genes can be detected  Hendrickson et al. (2002); Carreon-Diazconti et al. (2009); Matturro et al. (2013a, 2013b); Hug and Edwards (2013)  FISH  Direct  Fluorescent in situ hybridisation, microscopic detection of labelled microorganisms; possibility to label specific groups of microorganisms and/or genes  Direct visualisation of microorganisms present  Challenging for Dehalococcoides spp. due to low amount of target  Matturro et al. (2012, 2013a, 2013b, 2016); Matturro and Rossetti (2015)  Metagenomics  Indirect  Sequencing of the extractable DNA from the bacterial community of the site  Primer-independent detection/relative quantification of known dehalogenating bacteria and resp. enzymes  Handling and processing of large datasets  Reiss, Guerra and Makhnin (2016); Weigold et al. (2016)  Metaproteomics  Direct  Analysis of enzymes present  Direct evidence of functionality  No methods available/published so far for complex natural environments  None  Tracer experiments  Application of in situ stable isotope tracers (BACTRAP)  Direct  Detection of the incorporation of a stable isotope label (13C) into biomass or metabolites  Direct evidence for use as growth substrate (analysis biomass), pathways (metabolites, protein-SIP), identification of microbial community involved (DNA/RNA SIP)  Only usable if the contaminant is used as the carbon source; still under development  Kittelmann and Friedrich (2008a,b); Stelzer et al. (2006)  Ex situ  Laboratory microcosms with spiked organohalides  Direct evidence for the potential  Dehalogenation products  Direct evidence for the potential capacity of the in situ community  Long time frame; change of environment (laboratory vs. in situ)  Fennell et al. (2001); Fagervold et al. (2006); Fung et al. (2009)  Approach/ concept  Direct/ indirect proof  Indicators and principle  Advantage  Challenges  Examples/references  Hydrochemistry  Redox processes  Indirect  Presence of methane, sulphide and reduced iron (FeII), low redox and absence of oxygen indicate anoxic conditions required for reductive dehalogenation  Easy to analyse (iron, sulphate, nitrate etc.)  Process may be linked to other substrates  Nijenhuis et al. (2007a, 2009, 2013)  Contamination pattern  Direct  Concentration decrease may be a result of degradation but also dilution or sorption  Easy to analyse (GC-FID/MS; HPLC, etc.)  Co-contaminations, decrease of concentration due to sorption, dilution, etc.  Heidrich, Weiß and Kaschl (2004b)  Metabolite detection  Direct  Detection of dehalogenation products such as TCE, DCE, VC and ethene  Proof for pathway  Metabolite may be present as co-contaminant or as degradation product from multiple processes  Heidrich, Weiß and Kaschl (2004b)  Compound-specific stable isotope analysis  Direct  Enrichment of heavy stable isotopes (13C, 2H, 37Cl) indicates degradation of the compound; in reductive dehalogenation correlated with a depleted signature in the final product  Qualitative and potentially quantitative assessment of biodegradation  Equipment not universally accessible and applicable; complementary methods needed for quantitative assessment; reference laboratory studies not always available; isotope compositions of intermediates are difficult to interpret  Kaschl et al. (2005); Imfeld et al. (2008b); Nijenhuis et al. (2016)  Biomarker/molecular biology  PCR/qPCR  Direct  rRNA and rRNA genes of OHRB  Rapid identification of organisms present  Not or only an indirect indicator of activity; only known organisms and genes can be detected  Hendrickson et al. (2002); Carreon-Diazconti et al. (2009); Matturro et al. (2013a, 2013b); Hug and Edwards (2013)  FISH  Direct  Fluorescent in situ hybridisation, microscopic detection of labelled microorganisms; possibility to label specific groups of microorganisms and/or genes  Direct visualisation of microorganisms present  Challenging for Dehalococcoides spp. due to low amount of target  Matturro et al. (2012, 2013a, 2013b, 2016); Matturro and Rossetti (2015)  Metagenomics  Indirect  Sequencing of the extractable DNA from the bacterial community of the site  Primer-independent detection/relative quantification of known dehalogenating bacteria and resp. enzymes  Handling and processing of large datasets  Reiss, Guerra and Makhnin (2016); Weigold et al. (2016)  Metaproteomics  Direct  Analysis of enzymes present  Direct evidence of functionality  No methods available/published so far for complex natural environments  None  Tracer experiments  Application of in situ stable isotope tracers (BACTRAP)  Direct  Detection of the incorporation of a stable isotope label (13C) into biomass or metabolites  Direct evidence for use as growth substrate (analysis biomass), pathways (metabolites, protein-SIP), identification of microbial community involved (DNA/RNA SIP)  Only usable if the contaminant is used as the carbon source; still under development  Kittelmann and Friedrich (2008a,b); Stelzer et al. (2006)  Ex situ  Laboratory microcosms with spiked organohalides  Direct evidence for the potential  Dehalogenation products  Direct evidence for the potential capacity of the in situ community  Long time frame; change of environment (laboratory vs. in situ)  Fennell et al. (2001); Fagervold et al. (2006); Fung et al. (2009)  View Large Within Europe, chlorinated hydrocarbons are among the most frequently detected contaminants, and are found at 10% of a total of around 340 000 sites that are likely to require remediation (van Liedekerke et al. 2014). One such contaminated field site is the Bitterfeld-Wolfen megasite (Fig. 1). Although too extensive for active remediation, this field site has been subject to numerous studies over the last two decades related to the processes contributing to organohalide removal, mainly of chlorinated ethenes, monochlorobenzene and chlorinated dioxins. The aim of this review is to summarise the current state of the research and knowledge related to this region with a focus on the observed attenuation reactions of organochlorines in situ, the enriched and isolated Dehalococcoides mccartyi strains derived from this region and their ecophysiology towards site-specific pollutants. Figure 1. View largeDownload slide Map of the Bitterfeld-Wolfen region with its location in Germany (inset) indicating the spatial dimension of the environmental contamination with groundwater plume hotspots and investigated areas using microbiological approaches: chlorinated ethenes, around well BVV3051 from which strain BTF08 was enriched and isolated; MCB around SAFIRA lab area; dioxins around Spittelwasser from the sediments of which strain DCMB5 was isolated; DCE around Bergmannshof. Figure 1. View largeDownload slide Map of the Bitterfeld-Wolfen region with its location in Germany (inset) indicating the spatial dimension of the environmental contamination with groundwater plume hotspots and investigated areas using microbiological approaches: chlorinated ethenes, around well BVV3051 from which strain BTF08 was enriched and isolated; MCB around SAFIRA lab area; dioxins around Spittelwasser from the sediments of which strain DCMB5 was isolated; DCE around Bergmannshof. THE BITTERFELD-WOLFEN MEGASITE: HISTORY AND CONTAMINATION Groundwater contamination About 100 years of intensive chlorine-based chemical industry and neighbouring open-pit lignite mining left a large-scale soil and groundwater contamination in a dimension rarely seen previously (Gossel, Stollberg and Wycisk 2009). Nearby lignite deposits as primary input for the carbon-based chemistry and local energy production, potash mining, water resources from adjacent river streams as well as available cheap labour provided favourable site conditions for the rise of the Bitterfeld-Wolfen region into one of the world's largest chemical industrial centres at the beginning of the 20th century. Electrochemical industry started the synthesis of caustic soda and chlorine lime in 1893 and was rapidly expanded by complementary production sectors of e.g. hydrogen, chlorate, liquid chlorine and chlorobenzene until 1910. At the beginning of World War I in 1914, the chemical industry was also manufacturing explosives and chemical warfare agents. As a consequence of the development of innovative products such as the world's first fabrication of polyvinyl chloride, as well as the output of powdered metals, alloys, industrial cleaners, cellulose, synthetic fibres, pesticides, ion exchangers and many other direct consumer goods, the region had grown to a major of international producer of industrial chemicals. By 1969, its product range covered approximately 4500 individual chemical compounds and intensive chemical industry continued there until Germany’s reunification in 1989/90 (Stollberg 2013). Over a period of approximately a century, the intensive long-term chemical production coupled with various local disasters, inevitable handling losses and poor waste management, resulted in the ‘uncontrolled’ release and dumping of residuals from chemical industry. Combined, these events caused a multisource soil and groundwater contamination in an area of about 25–30 km2 (Wycisk 2003), with an affected groundwater volume of >200 000 000 m3 (Dermietzel and Christoph 2001; Heidrich et al. 2004). All environmental compartments such as air (Popp et al. 2000), soil and surface waters (Rückert et al. 2005) were affected. Especially, the groundwater system was heavily polluted by benzene, toluene, ethylbenzene and xylenes (BTEX), chlorobenzene, volatile organic compounds (VOCs), including tetra- and trichloroethene (PCE and TCE), hexachlorocyclohexanes (HCHs), phenols, heavy metals and other organic contaminants. Synthesis of PCE and TCE had started in 1925, and these compounds have been used extensively as degreasing, cleansing and extraction agents in the local metalworking, glass, optical and textile industries (Fischer 2004). As a consequence, this area has been the object of intensive interdisciplinary research since around 1997. Regional foci of severe pollutions include the following: (i) the contamination hot spot at the former chlorobenzene synthesis plant within the industrial area (see Fig. 1); (ii) a widespread groundwater contamination VOC plume beneath Bitterfeld's urban area that is heading downgradient towards the Mulde River and nearby surface waters (Kaschl et al. 2005; Imfeld et al. 2011); and (iii) the Spittelwasser floodplain north of Bitterfeld-Wolfen, which, according to Wycisk et al. (2013), has been shown and remains a distinct source area for HCH mobilisation into the Elbe River (Fig. 1). A major focus of environmental research and accompanying technical remediation measures was to tackle the affected regional groundwater system (Kaschl, Rügner and Weiß 2004; Heidrich, Weiß, Kaschl 2004). This aquifer system consists of a lower aquifer (tertiary marine fine sands) and upper aquifer (quaternary glacio-fluvial sands and gravels). Both are separated from each other by the regionally distributed lignite seam complex, an aquitard formed by various tertiary lignite seams and silts and clays. The lignite seam complex is only absent in areas of geogenic sub-glacial or fluvial erosion and because of former open-pit lignite mining activities. Mining-related dewatering of the upper and part of the lower aquifer induced aerobic conditions, and this caused the weathering of sulfidic minerals such as pyrite or marcasite, which are associated with sediments of the lignite seam complex. As a consequence of abandoned mining activities and associated water abstractions, a groundwater rebound followed and subsequently caused hydrochemical variations related to acid-mine drainage. This mechanism affects the surrounding groundwater and is characterised by lowering of pH and increased loads of Fe2+ and SO42−. Moreover, high clay contents of lignite-containing sediments, the lignite itself as well as the increased fraction of organic carbon in aquifer sediments resulted in high contaminant-specific retardation and sorption rates of the underlying long-term solute transport beneath the Bitterfeld-Wolfen area (Dermietzel and Christoph 2001). Chlorinated aromatics in Spittelwasser sediment The heavy contamination of Spittelwasser sediments with polychlorinated dibenzo-p-dioxins and -furans and HCH isomers has attracted considerable public attention. This is principally because sediment contaminants mobilised during high flood events, settle further downstream along the Mulde-Elbe river system all the way to the North Sea, strongly impairing the agricultural use of the flood plains (Götz and Lauer 2003; Götz et al. 2007). More than 10 years after the discharge of uncleaned industrial wastewater into the Spittelwasser tributary ceased, the load of sediments with chlorinated PCDD/Fs, HCHs, chlorinated benzenes and polychlorinated naphthalenes was still in the mg per kg range (Brack et al. 2003; Bunge et al. 2007). These contaminants may have persisted since decades dating back to their initial production, e.g. hexachlorobenzene has been produced since 1896. The PCDD/F congener pattern and specific fingerprints of tetra- and penta-chlorinated dibenzofurans pointed to a metallurgic process as the major source of the dioxin pollution. During World War II, between 1940 and 1945, the manufacturing of magnesium alloys was strongly intensified in Bitterfeld, based on the electrolysis of water-free MgCl2. The latter was produced in the so-called Bitterfeld process, which, from a current perspective, provided ideal conditions for the formation of PCDD/Fs: magnesium carbonates were mixed with brown coal and pitch and heated to 300°C–1000°C in a stream of chlorine gas (Büchen 1995). Waste waters from exhaust gas scrubbing were probably the main source of the high dioxin load in river sediments, which agrees with the abrupt increase in dioxin concentration in respective dated Elbe River sediment layers (Bunge et al. 2007; Götz et al. 2007). This long contamination history probably selected for a microbial community with a specific organohalide degradation potential. Microbial dehalogenation, in addition to natural perturbations in the dynamic river system, has most likely contributed to the subsequent changes in the pollutant profile. In this respect, it is interesting to note that a relatively high content of mono- to trichlorinated PCDD/F congeners has been identified in Spittelwasser sediment (Bunge et al. 2007). Biotransformation of organohalides Aerobic degradation of lower chlorinated ethenes and benzenes has been investigated and described extensively in literature, including the corresponding pathways (Field and Sierra-Alvarez 2008; Bradley and Chapelle 2010). Aerobic or coupled aerobic–anaerobic degradation was investigated specifically at the test site of the UFZ (SAFIRA project) and was found to be feasible in reactors amended with hydrogen peroxide (Alfreider, Vogt and Babel 2003; Balcke et al. 2004; Vogt et al. 2004). Furthermore, oxidation of dichloroethenes (DCEs) was indicated in a planted model wetland system (Imfeld et al. 2008a). The main focus of research in the Bitterfeld-Wolfen region was, however, on the anaerobic degradation activities due to the mainly anoxic conditions in the contaminated aquifers and sediments as indicated by the absence of oxygen and presence of reduced iron, sulphide and methane (Nijenhuis et al. 2007a; Imfeld et al. 2008b, 2011). Consequently, in the following sections we discuss the findings of research that has focused on anaerobic processes. Investigation of organohalide biotransformation in situ To get insight into the in situ dehalogenation process at the contaminated Bitterfeld site, different methods (a summary is shown in Table 1) were combined. The first approach to detect in situ activity was a detailed chemical analysis of the contaminants and their possible dehalogenation products. Compound-specific stable isotope analysis, detecting the enrichment of heavy stable isotopes (13C) in MCB, PCE and TCE and depletion in end products such as ethene, proved particularly useful to detect specific transformation activities in the Bitterfeld megasite (e.g. Kaschl et al. 2005; Imfeld et al. 2008b; Nijenhuis et al. 2016). Cultivation-independent methods such as PCR were used to detect the presence of specific microorganisms. Briefly, DNA was extracted directly from the groundwater sample and PCR with genus-specific primers targeting the 16S rRNA genes of well-known reductively dehalogenating bacteria such as Dehalococcoides, Dehalobacter and Desulfitobacterium was applied. In addition, genes encoding functionally described reductive dehalogenases were also amplified (e.g. Nijenhuis et al. 2007a; Imfeld et al. 2010; Mészáros et al. 2013). Both results are a clear indication for the dehalogenation potential at the respective site. However, the responsibility of the identified organisms for the observed degradation process cannot be undoubtedly clarified, because only known organisms and genes are detected. State-of-the-art activity-directed approaches such as transcriptomics or metaproteomics are still hard to realise in case of poorly colonised environments. Therefore, in most cases, laboratory microcosms and subsequent enrichment cultures were set up. Sometimes, pure cultures could be obtained, which showed the same dehalogenation/degradation pathway as observed in situ (Bunge et al. 2008; Kaufhold et al. 2013). Thus, the involvement of specific organisms could be indirectly confirmed. Tracer experiments, using 13C-labeled compounds, combined with DNA-stable isotope probing were another option, which enables the non-targeted detection of potentially involved organisms (Martinez-Lavanchy et al. 2011). In the following sections, the different approaches and challenges to uncover the ecophysiology of organohalide-degrading bacteria at the Bitterfeld site will be summarised, structured by compound class. Dehalogenation of chloroethenes in groundwater The most extensive investigations have been carried out with chlorinated ethenes, focusing on an area of approximately 2 km2 that includes the groundwater well BVV3051 and ‘Bergmannshof’ (Fig. 1). Early investigations already indicated the reductive dechlorination of chlorinated ethenes in situ because vinyl chloride (VC) and DCEs were detected in addition to PCE and TCE (Heidrich, Weiß, Kaschl 2004). Ethene, as final product of the dehalogenation, was not analysed initially, but later investigations showed the presence of this metabolite (Nijenhuis et al. 2007a). Compound-specific stable carbon isotope analysis supported in situ reductive dehalogenation to ethene as a main process contributing to chloroethene detoxification at the field site, with relatively depleted ethene isotope signatures compared to precursors such as PCE or DCEs (Nijenhuis et al. 2007a; Imfeld et al. 2008b, 2011). Laboratory microcosms confirmed the potential for reductive dehalogenation by the native microbial community as PCE was fully dehalogenated to ethene (Nijenhuis et al. 2007a). The dechlorination of both cis- and trans-DCE was observed in a model-constructed wetland running with groundwater from the area in ‘Bergmannshof’ (Imfeld et al. 2008a). Furthermore, the search for OHRB indicated the consistent presence of Dehalococcoides spp. in the investigated industrial area at the field site, with a less frequent occurrence of Dehalobacter spp., Geobacter and Desulfuromonas spp. This suggested a main contribution of Dehalococcoides spp. towards the observed dehalogenation activity (Nijenhuis et al. 2007a; Imfeld et al. 2008b, 2010, 2011). Throughout the tertiary and quarternary aquifers, even down to ∼50 m below the surface, chemical and molecular biological markers supported the presence of organohalide respiration and the corresponding bacteria (Imfeld et al. 2011). Analysing a subset of samples, a homogeneous Dehalococcoides sp. affiliated with the Pinellas subgroup was observed in groundwater samples as well as laboratory systems, including microcosms and wetland derived from the groundwater from the edge of the DCE plume at ‘Bergmannshof’ (Mészáros et al. 2013). Interestingly, tceA, coding for the TCE reductive dehalogenase, could not be detected in any of the samples, even though D. mccartyi strain BTF08 was enriched from samples collected from the same site. Strain BTF08 is capable of dechlorinating TCE and possesses the tceA gene (Mészáros et al. 2013; Pöritz et al. 2013). Variable bvcA sequences were detected only in enriched systems such as microcosms and constructed wetlands, while identical vcrA sequences were detected in all types of samples (Mészáros et al. 2013). Furthermore, 12 additional groundwater wells from distant locations within the Bitterfeld megasite exhibiting different patterns of contamination by chlorinated ethenes and benzenes were examined for the presence of known OHRB. Nine wells contained D. mccartyi strains, some of which also contained Dehalobacter and/or Desulfitobacterium based on 16S rRNA gene sequences and all contained at least one of five tested rdhA genes, suggesting active reductive dechlorination was occurring. The three remaining wells were characterised by pH values below 5.5 or a redox potential > +330 mV, pointing to the limits of the natural attenuation at the Bitterfeld site (unpublished data). Chlorobenzene degradation in groundwater Chlorobenzenes, mainly monochlorobenzene, are another relevant contaminant in the Bitterfeld-Wolfen region (Heidrich et al. 2004). Monochlorobenzene may be present as a result of contamination but also could result from dehalogenation of higher chlorinated benzenes or hexachlorocyclohexanes (Field and Sierra-Alvarez 2008; Lal et al. 2010). Dehalococcoides mccartyi strain BTF08, which was enriched from the Bitterfeld groundwater (see below), was observed to reduce γ-HCH to MCB and benzene (Kaufhold et al. 2013), suggesting that this strain could contribute to MCB production in situ. One main anoxic MCB plume was investigated in more detail (see Fig. 1), and in this case a decrease in concentration was associated with an enrichment of 13C in MCB, indicating anaerobic biodegradation (Kaschl et al. 2005). Further studies involving [13C6]-labelled MCB confirmed the utilisation of MCB as a carbon source by indigenous microorganisms under anoxic conditions in both in situ microcosms and laboratory microcosms (Nijenhuis et al. 2007b). The microorganisms and their corresponding pathways have not been resolved yet. Iron oxides, however, were observed to stimulate mineralisation of MCB (Schmidt et al. 2014). In separate experiments, applying [13C6]-MCB as a tracer, DNA-stable isotope probing in combination with the sequencing of bands from single-strand conformation polymorphism analysis allowed to investigate the MCB-metabolising community. This was observed to include bacteria from the phyla Proteobacteria, Fibrobacteres as well as from the candidate division OD1 (Martinez-Lavanchy et al. 2011). To date, further enrichment attempts have failed. Dioxin-dechlorination potential in contaminated freshwater sediments Sediments of different creeks and rivers in the area East and North of Bitterfeld, which are subject to annual high floods, are highly polluted with PCDD/Fs. Total concentrations of 3400 and 7300 ng (kg d.w.)−1 in the Leine-Durchstich creek and the river Mulde (unpublished data), respectively, and 8560 000 ng (kg d.w.)−1 for the Spittelwasser creek (Bunge et al. 2007) were detected. The potential of the intrinsic microbial community to dechlorinate PCDDs was demonstrated with microcosms spiked with 1,2,3,4-tetrachlorodibenzo-p-dioxin (TeCDD) as a model congener (Bunge and Lechner 2001; Bunge, Ballerstedt and Lechner 2001). Lower chlorinated products were detected in microcosms from all sources, including the sulfate-rich acidic (pH 4) Leine-Durchstich sediment (Fig. 1) and six samples from Spittelwasser (representing two neighbouring sampling sites at three different depths). The final dechlorination products were 1,3- and 2,3-dichlorodibenzo-p-dioxin in all cases. Three dechlorination pathways, characterised by different combinations of chlorine removal from lateral and peripheral positions (Fig. 2), were identified in subcultures supplemented with the possible intermediates 1,2,3- and 1,2,4-trichlorodibenzo-p-dioxin (TrCDD), suggesting the presence of diverse dioxin-dechlorinating communities (Bunge and Lechner 2001; Bunge, Ballerstedt and Lechner 2001). Dehalococcoides mccartyi and Desulfitobacterium spp. were detected in these subcultures. In analogy to the dioxin-dechlorinating D. mccartyi strain CBDB1, which was shown to dechlorinate selected dioxin congeners (Bunge et al. 2003), a role of the sediment-derived D. mccartyi population in PCDD dechlorination was suggested and further supported by its growth during dechlorination of 1,2,3- and 1,2,4-TrCDD (Ewald et al. 2007). Figure 2. View largeDownload slide Pathways for the reductive dechlorination of the model congener 1,2,3,4-TeCDD. Sediment microcosms from the Mulde river and the creeks Spittelwasser and Leine-Durchstich dechlorinated 1,2,3,4-TeCDD by different sequences of chlorine removal from peripheral (peri) and lateral positions (solid arrows) resulting in three different patterns of products (see e.g. Adrian and Lechner 2004). Dehalococcoides mccartyi strain DCMB5 isolated from Spittelwasser showed a pathway restricted to the removal of peripheral chlorines only (empty arrows), suggesting that the sediments harbour a much larger diversity of dioxin-dechlorinating bacteria. TrCDD, DCDD, MCDD: tri- di- and monochlorodibenzo-p-dioxin. Figure 2. View largeDownload slide Pathways for the reductive dechlorination of the model congener 1,2,3,4-TeCDD. Sediment microcosms from the Mulde river and the creeks Spittelwasser and Leine-Durchstich dechlorinated 1,2,3,4-TeCDD by different sequences of chlorine removal from peripheral (peri) and lateral positions (solid arrows) resulting in three different patterns of products (see e.g. Adrian and Lechner 2004). Dehalococcoides mccartyi strain DCMB5 isolated from Spittelwasser showed a pathway restricted to the removal of peripheral chlorines only (empty arrows), suggesting that the sediments harbour a much larger diversity of dioxin-dechlorinating bacteria. TrCDD, DCDD, MCDD: tri- di- and monochlorodibenzo-p-dioxin. ENRICHMENT OF D. MCCARTYI STRAINS FROM THE BITTERFELD SITE Dehalococcoides mccartyi strain BTF08: a dedicated chloroethene-respiring groundwater isolate Isolation of microorganisms relevant for the observed activity at a field site is a challenge and associated with many artefacts because laboratory conditions may favour growth of minor populations. Interestingly, a homogeneous D. mccartyi population based on 16S rRNA gene sequence was observed in Bitterfeld groundwater as well as enrichment cultures (Mészáros et al. 2013). Initially, enrichments from groundwater of well BVV3051 contaminated mainly with DCEs were prepared and shown to be capable of the complete reductive dechlorination of tetrachloroethene to ethene with lactate as electron donor (Nijenhuis et al. 2007a). Subsequent transfers, initially with lactate as electron donor and carbon source, later with hydrogen and acetate as electron donor and carbon source, respectively, resulted in a culture dominated by a Dehalococcoides sp. (Cichocka et al. 2010). The novel strain, D. mccartyi strain BTF08, can couple all dechlorination steps of PCE to ethene to energy conservation (Cichocka et al. 2010) and, accordingly, its genome encodes all three necessary enzymes (PceA, TceA and VcrA) (Pöritz et al. 2013). In addition to the chlorinated ethenes, 1,2-dichloroethane, higher chlorinated benzenes and γ-hexachlorocyclohexane were also reductively dechlorinated by this strain (Fig. 3) (Kaufhold et al. 2013). Overall, strain BTF08 appears to be more specialised toward chlorinated alkanes and alkenes and only dechlorinates highly chlorinated benzenes to trichlorobenzenes (Cichocka et al. 2010; Kaufhold et al. 2013; Schmidt, Lege and Nijenhuis 2014). Figure 3. View largeDownload slide Pathways of the reductive dechlorination of chlorinated ethenes and 1,2-dichloroethane (DCA) by D. mccartyi strain BTF08 (solid arrows) and strain DCMB5 (open arrow). The dashed arrow indicates the minor amounts of VC produced during reductive dechlorination of 1,2-DCA by strain BTF08. Figure 3. View largeDownload slide Pathways of the reductive dechlorination of chlorinated ethenes and 1,2-dichloroethane (DCA) by D. mccartyi strain BTF08 (solid arrows) and strain DCMB5 (open arrow). The dashed arrow indicates the minor amounts of VC produced during reductive dechlorination of 1,2-DCA by strain BTF08. Dehalococcoides mccartyi strain DCMB5: a dioxin-dechlorinating isolate from Spittelwasser sediment In an attempt to isolate a dioxin-dechlorinating bacterium from Spittelwasser sediment, 1,2,3-trichlorobenzene (1,2,3-TrCB) was added as an alternative electron acceptor to cultures previously sub-cultured on 1,2,4-TrCDD. 1,2,3-TrCB was applied in a hexadecane phase at high concentrations. It was rapidly dechlorinated to 1,3-dichlorobenzene. A strong increase of bacteria capable to dechlorinate 1,2,4-TrCDD was observed (Bunge et al. 2008). Successive inhibition of methanogens and gram-positive bacteria increased the number of Dehalococcoides cells and finally allowed the isolation of strain DCMB5 from agar shake dilutions (Bunge et al. 2008). This strain showed a strong preference for the dechlorination of peripheral chlorine substituents (Fig. 2) from dioxin congeners and even removed the single chlorine from 1-monochlorodibenzo-p-dioxin (Pöritz et al. 2015), demonstrating its potential contribution to the complete dechlorination of dioxin congener mixtures. Moreover, it was capable of dechlorinating a broad range of chlorinated benzenes and phenols as well as PCE, which it dechlorinated to TCE (Fig. 3). The preference for the dechlorination of chlorinated aromatics might be the result of the enrichment procedure, but it also reflects its origin from heavily dioxin- and chlorobenzene-polluted sediment. Accordingly, its genome does not encode any classical dehalogenase responsible for the dechlorination of chlorinated ethenes such as PceA, TceA or VcrA. Rather, it encodes an ortholog of CbrA, the chlorobenzene reductive dehalogenase in strain CBDB1 (Adrian et al. 2007; Pöritz et al. 2013). As expected, this ortholog, Dcmb_86, was the most abundant dehalogenase in proteomes of pentachlorobenzene- and 1,2,3-TrCB-grown cells, but also of PCE-grown cells, suggesting a broader substrate specificity (Pöritz et al. 2015). Another property of strain DCMB5 is the capability to form type IV pili, detected by transmission electron microscopy of exponentially growing cells. The corresponding gene cluster is conserved in D. mccartyi suggesting its significance for its natural lifestyle. CONCLUSIONS Despite the large horizontal and vertical dimensions of an extremely complex multisource environmental pollution, first insights into an ongoing natural attenuation remediation strategy and related processes at the Bitterfeld megasite were obtained. Based on a relatively dense monitoring network of sampling wells, thorough chemical and hydrogeological analyses (Wycisk et al. 2013), and the selection of site-specific contamination profiles, biomarkers, microbiological and stable isotope analyses were applied and throughout revealed the presence of actively OHRB (see also Table 1). It is interesting to note that D. mccartyi was present at sites contaminated with chloroethenes despite the high concentrations of co-contaminating benzenes and BTEX (Heidrich, Weiß, Kaschl 2004; Nijenhuis et al. 2007a; Cichocka et al. 2010). This suggests that D. mccartyi is relatively resistant to the presence of hydrophobic, non-halogenated contaminants, whereas it is more dependent on suitable ranges of pH and redox potential. This indicates that the natural attenuation of chlorinated ethenes is ongoing at several places within the Bitterfeld megasite and was particularly shown for the DCE plume near the Bergmannshof area. Strain BTF08 isolated from well BVV3051 can be used as an example to simulate the influence of groundwater parameters on the dechlorination process. Its properties reflect well the ecophysiological adaptation of D. mccartyi to this site. The presence of the vcrA gene enabled efficient dechlorination of the main contaminants DCE and VC. It was also able to dechlorinate PCE and TCE in accord with the presence of the respective pceA and tceA genes. The tceA gene was not detected in situ and, thus, strain BTF08 might represent a minor population at the field site (Mészáros et al. 2013). Even though, PCE and TCE were dechlorinated throughout the field site (Imfeld et al. 2011). Overall, D. mccartyi appears to be more capable to adapt to a variety of environmental conditions compared to e.g. Dehalobacter, Geobacter or Desulfuromonas spp., as D. mccartyi was more frequently detected in groundwater samples. For monochlorobenzene, which is one of the organic pollutants with the highest groundwater concentrations in Bitterfeld (>50 000 μg/l at the former chlorobenzene synthesis plant, Wycisk 2003; Kaschl et al. 2005), continuous natural attenuation processes are indicated and have been verified; however, the organisms responsible are not yet known. Although the dioxin-dechlorinating potential was demonstrated for the heavily contaminated Spittelwasser sediment and the isolate D. mccartyi DCMB5, it must be assumed that the natural attenuation process might be slow due to the low water solubility and strong sorptive behaviour of PCCD/Fs. The in situ assessment of PCDD/F dechlorination, e.g. by carbon isotope fractionation analyses, is a difficult task due to the complex congener patterns and the difficulty in extracting sufficient amounts of the compounds. The lower chlorinated products might, however, be a suitable target for analysis (Ewald et al. 2007; Liu et al. 2010). The hexachlorocyclohexane isomers are another problematic contaminant group at the Bitterfeld site, where large deposits represent secondary sources for continuing contamination of groundwater and surface water. Although initial data showed reductive dechlorination of γ-HCH by D. mccartyi strain BTF08 (Kaufhold et al. 2013), and isotopic and enantiomer patterns suggest in situ biodegradation of HCHs in groundwater at the Bitterfeld site (Liu et al. 2017), there is still a gap in our understanding of the intrinsic potential of the microbial communities in groundwater and river sediments and flood plains for the attenuation of HCH isomers, which will need to be addressed in the future. To further address the activity and provide an insight into the ecophysiology of anaerobic microbial dehalogenating bacteria, a suite of methods is already available (Table 1). Currently, however, mainly a direct indicator of activity, such as could be achieved by direct detection of the reductive dehalogenases, via proteomic approaches, is lacking. A combination of methods, such as the analysis of geochemical conditions, substrate and metabolite analysis, compound-specific stable isotope analysis in combination with specific detection of dehalogenating bacteria and their functional genes and a statistical framework, however, may provide such initial insights (Atashgahi et al. 2017). Acknowledgements We would like to thank Marlén Pöritz for sharing unpublished data and Gary Sawers for English proofreading. FUNDING This study was funded by the Deutsche Forschungsgemeinschaft (Research Unit FOR 1530, NI 1329/1-1 and/1-2 as well as LE 780/4-1 and/4-2) and the Helmholtz Centre for Environmental Research—UFZ. Conflict of interest. None declared. REFERENCES Adrian L, Lechner U. Reductive dechlorination of polychlorinated benzenes and dioxins. 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Anaerobic microbial dehalogenation and its key players in the contaminated Bitterfeld-Wolfen megasite

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Abstract

Abstract The megasite Bitterfeld-Wolfen is highly contaminated as a result of accidents and because of dumping of wastes from local chemical industries in the last century. A variety of contaminants including chlorinated ethenes and benzenes, hexachlorohexanes and chlorinated dioxins can still be found in the groundwater and (river) sediments. Investigations of the in situ microbial transformation of organohalides have been performed only over the last two decades at this megasite. In this review, we summarise the research on the activity of anaerobic dehalogenating bacteria at the field site in Bitterfeld-Wolfen, focusing on chlorinated ethenes, monochlorobenzene and chlorinated dioxins. Various methods and concepts were applied including ex situ cultivation and isolation, and in situ analysis of hydrochemical parameters, compound-specific stable isotope analysis of contaminants, 13C-tracer studies and molecular markers. Overall, biotransformation of organohalides is ongoing at the field site and Dehalococcoides mccartyi species play an important role in the detoxification process in the Bitterfeld-Wolfen region. in situ, organohalide biotransformation, reductive dechlorination, Dehalococcoides, monochlorobenzene, chlorinated ethenes, chlorinated dibenzo-p-dioxins INTRODUCTION Since the discovery of reductive dechlorination as a respiratory process (Dolfing and Tiedje 1987; Dolfing 1990) and its potential as a remediation strategy for organohalides (Holliger et al. 1997), research has extended from the investigation of the microorganisms and enzymes involved to the identification of these microorganisms and the responsible genes at contaminated field sites. Bioremediation and bioaugmentation strategies to clean up contaminated sites, especially those with chlorinated ethenes, have been implemented and the detection of microorganisms, such as Dehalococcoides, and their reductive dehalogenase genes (rdhA) has become routine (Major et al. 2002; Ritalahti et al. 2005; Stroo, Major and Gossett 2010). Several methods are currently available, usually applied in combination (for an overview see Table 1), for the investigation of the ecophysiology of microorganisms capable of the degradation of organohalides under anoxic conditions. For site assessment, more traditional methods such as analysis of the geochemistry as an indicator of the geochemical conditions, and contaminant and product concentrations as indicators for (bio)transformation are used in combination with newer approaches. These new approaches include compound-specific stable isotope analysis (Hunkeler et al. 2008; Nijenhuis et al. 2016) and detection of biomarkers, i.e. 16S rRNA genes of organohalide-respiring bacteria (OHRB) and their functional genes (Wilson 2010; Nijenhuis and Kuntze 2016). Combined ex situ, laboratory and microcosm approaches can be applied to investigate the potential for biotransformation of a contaminant and elucidation of conditions that improve activity. Stable isotope tracer approaches, applying a 13C-labelled contaminant in in situ microcosms, have been implemented in cases where the contaminant may be used as a carbon source and can be traced into metabolites and microbial biomass (Bombach et al. 2010). This method does not allow direct detection of OHRB as the carbon present in the contaminant is not used as carbon source but an external carbon source, e.g. acetate, is needed (Stelzer et al. 2006; Kittelmann and Friedrich 2008b). Table 1. Indicators/methods applied for the investigation of organohalide degradation in situ (adapted from Bombach et al. 2010). Approach/ concept  Direct/ indirect proof  Indicators and principle  Advantage  Challenges  Examples/references  Hydrochemistry  Redox processes  Indirect  Presence of methane, sulphide and reduced iron (FeII), low redox and absence of oxygen indicate anoxic conditions required for reductive dehalogenation  Easy to analyse (iron, sulphate, nitrate etc.)  Process may be linked to other substrates  Nijenhuis et al. (2007a, 2009, 2013)  Contamination pattern  Direct  Concentration decrease may be a result of degradation but also dilution or sorption  Easy to analyse (GC-FID/MS; HPLC, etc.)  Co-contaminations, decrease of concentration due to sorption, dilution, etc.  Heidrich, Weiß and Kaschl (2004b)  Metabolite detection  Direct  Detection of dehalogenation products such as TCE, DCE, VC and ethene  Proof for pathway  Metabolite may be present as co-contaminant or as degradation product from multiple processes  Heidrich, Weiß and Kaschl (2004b)  Compound-specific stable isotope analysis  Direct  Enrichment of heavy stable isotopes (13C, 2H, 37Cl) indicates degradation of the compound; in reductive dehalogenation correlated with a depleted signature in the final product  Qualitative and potentially quantitative assessment of biodegradation  Equipment not universally accessible and applicable; complementary methods needed for quantitative assessment; reference laboratory studies not always available; isotope compositions of intermediates are difficult to interpret  Kaschl et al. (2005); Imfeld et al. (2008b); Nijenhuis et al. (2016)  Biomarker/molecular biology  PCR/qPCR  Direct  rRNA and rRNA genes of OHRB  Rapid identification of organisms present  Not or only an indirect indicator of activity; only known organisms and genes can be detected  Hendrickson et al. (2002); Carreon-Diazconti et al. (2009); Matturro et al. (2013a, 2013b); Hug and Edwards (2013)  FISH  Direct  Fluorescent in situ hybridisation, microscopic detection of labelled microorganisms; possibility to label specific groups of microorganisms and/or genes  Direct visualisation of microorganisms present  Challenging for Dehalococcoides spp. due to low amount of target  Matturro et al. (2012, 2013a, 2013b, 2016); Matturro and Rossetti (2015)  Metagenomics  Indirect  Sequencing of the extractable DNA from the bacterial community of the site  Primer-independent detection/relative quantification of known dehalogenating bacteria and resp. enzymes  Handling and processing of large datasets  Reiss, Guerra and Makhnin (2016); Weigold et al. (2016)  Metaproteomics  Direct  Analysis of enzymes present  Direct evidence of functionality  No methods available/published so far for complex natural environments  None  Tracer experiments  Application of in situ stable isotope tracers (BACTRAP)  Direct  Detection of the incorporation of a stable isotope label (13C) into biomass or metabolites  Direct evidence for use as growth substrate (analysis biomass), pathways (metabolites, protein-SIP), identification of microbial community involved (DNA/RNA SIP)  Only usable if the contaminant is used as the carbon source; still under development  Kittelmann and Friedrich (2008a,b); Stelzer et al. (2006)  Ex situ  Laboratory microcosms with spiked organohalides  Direct evidence for the potential  Dehalogenation products  Direct evidence for the potential capacity of the in situ community  Long time frame; change of environment (laboratory vs. in situ)  Fennell et al. (2001); Fagervold et al. (2006); Fung et al. (2009)  Approach/ concept  Direct/ indirect proof  Indicators and principle  Advantage  Challenges  Examples/references  Hydrochemistry  Redox processes  Indirect  Presence of methane, sulphide and reduced iron (FeII), low redox and absence of oxygen indicate anoxic conditions required for reductive dehalogenation  Easy to analyse (iron, sulphate, nitrate etc.)  Process may be linked to other substrates  Nijenhuis et al. (2007a, 2009, 2013)  Contamination pattern  Direct  Concentration decrease may be a result of degradation but also dilution or sorption  Easy to analyse (GC-FID/MS; HPLC, etc.)  Co-contaminations, decrease of concentration due to sorption, dilution, etc.  Heidrich, Weiß and Kaschl (2004b)  Metabolite detection  Direct  Detection of dehalogenation products such as TCE, DCE, VC and ethene  Proof for pathway  Metabolite may be present as co-contaminant or as degradation product from multiple processes  Heidrich, Weiß and Kaschl (2004b)  Compound-specific stable isotope analysis  Direct  Enrichment of heavy stable isotopes (13C, 2H, 37Cl) indicates degradation of the compound; in reductive dehalogenation correlated with a depleted signature in the final product  Qualitative and potentially quantitative assessment of biodegradation  Equipment not universally accessible and applicable; complementary methods needed for quantitative assessment; reference laboratory studies not always available; isotope compositions of intermediates are difficult to interpret  Kaschl et al. (2005); Imfeld et al. (2008b); Nijenhuis et al. (2016)  Biomarker/molecular biology  PCR/qPCR  Direct  rRNA and rRNA genes of OHRB  Rapid identification of organisms present  Not or only an indirect indicator of activity; only known organisms and genes can be detected  Hendrickson et al. (2002); Carreon-Diazconti et al. (2009); Matturro et al. (2013a, 2013b); Hug and Edwards (2013)  FISH  Direct  Fluorescent in situ hybridisation, microscopic detection of labelled microorganisms; possibility to label specific groups of microorganisms and/or genes  Direct visualisation of microorganisms present  Challenging for Dehalococcoides spp. due to low amount of target  Matturro et al. (2012, 2013a, 2013b, 2016); Matturro and Rossetti (2015)  Metagenomics  Indirect  Sequencing of the extractable DNA from the bacterial community of the site  Primer-independent detection/relative quantification of known dehalogenating bacteria and resp. enzymes  Handling and processing of large datasets  Reiss, Guerra and Makhnin (2016); Weigold et al. (2016)  Metaproteomics  Direct  Analysis of enzymes present  Direct evidence of functionality  No methods available/published so far for complex natural environments  None  Tracer experiments  Application of in situ stable isotope tracers (BACTRAP)  Direct  Detection of the incorporation of a stable isotope label (13C) into biomass or metabolites  Direct evidence for use as growth substrate (analysis biomass), pathways (metabolites, protein-SIP), identification of microbial community involved (DNA/RNA SIP)  Only usable if the contaminant is used as the carbon source; still under development  Kittelmann and Friedrich (2008a,b); Stelzer et al. (2006)  Ex situ  Laboratory microcosms with spiked organohalides  Direct evidence for the potential  Dehalogenation products  Direct evidence for the potential capacity of the in situ community  Long time frame; change of environment (laboratory vs. in situ)  Fennell et al. (2001); Fagervold et al. (2006); Fung et al. (2009)  View Large Within Europe, chlorinated hydrocarbons are among the most frequently detected contaminants, and are found at 10% of a total of around 340 000 sites that are likely to require remediation (van Liedekerke et al. 2014). One such contaminated field site is the Bitterfeld-Wolfen megasite (Fig. 1). Although too extensive for active remediation, this field site has been subject to numerous studies over the last two decades related to the processes contributing to organohalide removal, mainly of chlorinated ethenes, monochlorobenzene and chlorinated dioxins. The aim of this review is to summarise the current state of the research and knowledge related to this region with a focus on the observed attenuation reactions of organochlorines in situ, the enriched and isolated Dehalococcoides mccartyi strains derived from this region and their ecophysiology towards site-specific pollutants. Figure 1. View largeDownload slide Map of the Bitterfeld-Wolfen region with its location in Germany (inset) indicating the spatial dimension of the environmental contamination with groundwater plume hotspots and investigated areas using microbiological approaches: chlorinated ethenes, around well BVV3051 from which strain BTF08 was enriched and isolated; MCB around SAFIRA lab area; dioxins around Spittelwasser from the sediments of which strain DCMB5 was isolated; DCE around Bergmannshof. Figure 1. View largeDownload slide Map of the Bitterfeld-Wolfen region with its location in Germany (inset) indicating the spatial dimension of the environmental contamination with groundwater plume hotspots and investigated areas using microbiological approaches: chlorinated ethenes, around well BVV3051 from which strain BTF08 was enriched and isolated; MCB around SAFIRA lab area; dioxins around Spittelwasser from the sediments of which strain DCMB5 was isolated; DCE around Bergmannshof. THE BITTERFELD-WOLFEN MEGASITE: HISTORY AND CONTAMINATION Groundwater contamination About 100 years of intensive chlorine-based chemical industry and neighbouring open-pit lignite mining left a large-scale soil and groundwater contamination in a dimension rarely seen previously (Gossel, Stollberg and Wycisk 2009). Nearby lignite deposits as primary input for the carbon-based chemistry and local energy production, potash mining, water resources from adjacent river streams as well as available cheap labour provided favourable site conditions for the rise of the Bitterfeld-Wolfen region into one of the world's largest chemical industrial centres at the beginning of the 20th century. Electrochemical industry started the synthesis of caustic soda and chlorine lime in 1893 and was rapidly expanded by complementary production sectors of e.g. hydrogen, chlorate, liquid chlorine and chlorobenzene until 1910. At the beginning of World War I in 1914, the chemical industry was also manufacturing explosives and chemical warfare agents. As a consequence of the development of innovative products such as the world's first fabrication of polyvinyl chloride, as well as the output of powdered metals, alloys, industrial cleaners, cellulose, synthetic fibres, pesticides, ion exchangers and many other direct consumer goods, the region had grown to a major of international producer of industrial chemicals. By 1969, its product range covered approximately 4500 individual chemical compounds and intensive chemical industry continued there until Germany’s reunification in 1989/90 (Stollberg 2013). Over a period of approximately a century, the intensive long-term chemical production coupled with various local disasters, inevitable handling losses and poor waste management, resulted in the ‘uncontrolled’ release and dumping of residuals from chemical industry. Combined, these events caused a multisource soil and groundwater contamination in an area of about 25–30 km2 (Wycisk 2003), with an affected groundwater volume of >200 000 000 m3 (Dermietzel and Christoph 2001; Heidrich et al. 2004). All environmental compartments such as air (Popp et al. 2000), soil and surface waters (Rückert et al. 2005) were affected. Especially, the groundwater system was heavily polluted by benzene, toluene, ethylbenzene and xylenes (BTEX), chlorobenzene, volatile organic compounds (VOCs), including tetra- and trichloroethene (PCE and TCE), hexachlorocyclohexanes (HCHs), phenols, heavy metals and other organic contaminants. Synthesis of PCE and TCE had started in 1925, and these compounds have been used extensively as degreasing, cleansing and extraction agents in the local metalworking, glass, optical and textile industries (Fischer 2004). As a consequence, this area has been the object of intensive interdisciplinary research since around 1997. Regional foci of severe pollutions include the following: (i) the contamination hot spot at the former chlorobenzene synthesis plant within the industrial area (see Fig. 1); (ii) a widespread groundwater contamination VOC plume beneath Bitterfeld's urban area that is heading downgradient towards the Mulde River and nearby surface waters (Kaschl et al. 2005; Imfeld et al. 2011); and (iii) the Spittelwasser floodplain north of Bitterfeld-Wolfen, which, according to Wycisk et al. (2013), has been shown and remains a distinct source area for HCH mobilisation into the Elbe River (Fig. 1). A major focus of environmental research and accompanying technical remediation measures was to tackle the affected regional groundwater system (Kaschl, Rügner and Weiß 2004; Heidrich, Weiß, Kaschl 2004). This aquifer system consists of a lower aquifer (tertiary marine fine sands) and upper aquifer (quaternary glacio-fluvial sands and gravels). Both are separated from each other by the regionally distributed lignite seam complex, an aquitard formed by various tertiary lignite seams and silts and clays. The lignite seam complex is only absent in areas of geogenic sub-glacial or fluvial erosion and because of former open-pit lignite mining activities. Mining-related dewatering of the upper and part of the lower aquifer induced aerobic conditions, and this caused the weathering of sulfidic minerals such as pyrite or marcasite, which are associated with sediments of the lignite seam complex. As a consequence of abandoned mining activities and associated water abstractions, a groundwater rebound followed and subsequently caused hydrochemical variations related to acid-mine drainage. This mechanism affects the surrounding groundwater and is characterised by lowering of pH and increased loads of Fe2+ and SO42−. Moreover, high clay contents of lignite-containing sediments, the lignite itself as well as the increased fraction of organic carbon in aquifer sediments resulted in high contaminant-specific retardation and sorption rates of the underlying long-term solute transport beneath the Bitterfeld-Wolfen area (Dermietzel and Christoph 2001). Chlorinated aromatics in Spittelwasser sediment The heavy contamination of Spittelwasser sediments with polychlorinated dibenzo-p-dioxins and -furans and HCH isomers has attracted considerable public attention. This is principally because sediment contaminants mobilised during high flood events, settle further downstream along the Mulde-Elbe river system all the way to the North Sea, strongly impairing the agricultural use of the flood plains (Götz and Lauer 2003; Götz et al. 2007). More than 10 years after the discharge of uncleaned industrial wastewater into the Spittelwasser tributary ceased, the load of sediments with chlorinated PCDD/Fs, HCHs, chlorinated benzenes and polychlorinated naphthalenes was still in the mg per kg range (Brack et al. 2003; Bunge et al. 2007). These contaminants may have persisted since decades dating back to their initial production, e.g. hexachlorobenzene has been produced since 1896. The PCDD/F congener pattern and specific fingerprints of tetra- and penta-chlorinated dibenzofurans pointed to a metallurgic process as the major source of the dioxin pollution. During World War II, between 1940 and 1945, the manufacturing of magnesium alloys was strongly intensified in Bitterfeld, based on the electrolysis of water-free MgCl2. The latter was produced in the so-called Bitterfeld process, which, from a current perspective, provided ideal conditions for the formation of PCDD/Fs: magnesium carbonates were mixed with brown coal and pitch and heated to 300°C–1000°C in a stream of chlorine gas (Büchen 1995). Waste waters from exhaust gas scrubbing were probably the main source of the high dioxin load in river sediments, which agrees with the abrupt increase in dioxin concentration in respective dated Elbe River sediment layers (Bunge et al. 2007; Götz et al. 2007). This long contamination history probably selected for a microbial community with a specific organohalide degradation potential. Microbial dehalogenation, in addition to natural perturbations in the dynamic river system, has most likely contributed to the subsequent changes in the pollutant profile. In this respect, it is interesting to note that a relatively high content of mono- to trichlorinated PCDD/F congeners has been identified in Spittelwasser sediment (Bunge et al. 2007). Biotransformation of organohalides Aerobic degradation of lower chlorinated ethenes and benzenes has been investigated and described extensively in literature, including the corresponding pathways (Field and Sierra-Alvarez 2008; Bradley and Chapelle 2010). Aerobic or coupled aerobic–anaerobic degradation was investigated specifically at the test site of the UFZ (SAFIRA project) and was found to be feasible in reactors amended with hydrogen peroxide (Alfreider, Vogt and Babel 2003; Balcke et al. 2004; Vogt et al. 2004). Furthermore, oxidation of dichloroethenes (DCEs) was indicated in a planted model wetland system (Imfeld et al. 2008a). The main focus of research in the Bitterfeld-Wolfen region was, however, on the anaerobic degradation activities due to the mainly anoxic conditions in the contaminated aquifers and sediments as indicated by the absence of oxygen and presence of reduced iron, sulphide and methane (Nijenhuis et al. 2007a; Imfeld et al. 2008b, 2011). Consequently, in the following sections we discuss the findings of research that has focused on anaerobic processes. Investigation of organohalide biotransformation in situ To get insight into the in situ dehalogenation process at the contaminated Bitterfeld site, different methods (a summary is shown in Table 1) were combined. The first approach to detect in situ activity was a detailed chemical analysis of the contaminants and their possible dehalogenation products. Compound-specific stable isotope analysis, detecting the enrichment of heavy stable isotopes (13C) in MCB, PCE and TCE and depletion in end products such as ethene, proved particularly useful to detect specific transformation activities in the Bitterfeld megasite (e.g. Kaschl et al. 2005; Imfeld et al. 2008b; Nijenhuis et al. 2016). Cultivation-independent methods such as PCR were used to detect the presence of specific microorganisms. Briefly, DNA was extracted directly from the groundwater sample and PCR with genus-specific primers targeting the 16S rRNA genes of well-known reductively dehalogenating bacteria such as Dehalococcoides, Dehalobacter and Desulfitobacterium was applied. In addition, genes encoding functionally described reductive dehalogenases were also amplified (e.g. Nijenhuis et al. 2007a; Imfeld et al. 2010; Mészáros et al. 2013). Both results are a clear indication for the dehalogenation potential at the respective site. However, the responsibility of the identified organisms for the observed degradation process cannot be undoubtedly clarified, because only known organisms and genes are detected. State-of-the-art activity-directed approaches such as transcriptomics or metaproteomics are still hard to realise in case of poorly colonised environments. Therefore, in most cases, laboratory microcosms and subsequent enrichment cultures were set up. Sometimes, pure cultures could be obtained, which showed the same dehalogenation/degradation pathway as observed in situ (Bunge et al. 2008; Kaufhold et al. 2013). Thus, the involvement of specific organisms could be indirectly confirmed. Tracer experiments, using 13C-labeled compounds, combined with DNA-stable isotope probing were another option, which enables the non-targeted detection of potentially involved organisms (Martinez-Lavanchy et al. 2011). In the following sections, the different approaches and challenges to uncover the ecophysiology of organohalide-degrading bacteria at the Bitterfeld site will be summarised, structured by compound class. Dehalogenation of chloroethenes in groundwater The most extensive investigations have been carried out with chlorinated ethenes, focusing on an area of approximately 2 km2 that includes the groundwater well BVV3051 and ‘Bergmannshof’ (Fig. 1). Early investigations already indicated the reductive dechlorination of chlorinated ethenes in situ because vinyl chloride (VC) and DCEs were detected in addition to PCE and TCE (Heidrich, Weiß, Kaschl 2004). Ethene, as final product of the dehalogenation, was not analysed initially, but later investigations showed the presence of this metabolite (Nijenhuis et al. 2007a). Compound-specific stable carbon isotope analysis supported in situ reductive dehalogenation to ethene as a main process contributing to chloroethene detoxification at the field site, with relatively depleted ethene isotope signatures compared to precursors such as PCE or DCEs (Nijenhuis et al. 2007a; Imfeld et al. 2008b, 2011). Laboratory microcosms confirmed the potential for reductive dehalogenation by the native microbial community as PCE was fully dehalogenated to ethene (Nijenhuis et al. 2007a). The dechlorination of both cis- and trans-DCE was observed in a model-constructed wetland running with groundwater from the area in ‘Bergmannshof’ (Imfeld et al. 2008a). Furthermore, the search for OHRB indicated the consistent presence of Dehalococcoides spp. in the investigated industrial area at the field site, with a less frequent occurrence of Dehalobacter spp., Geobacter and Desulfuromonas spp. This suggested a main contribution of Dehalococcoides spp. towards the observed dehalogenation activity (Nijenhuis et al. 2007a; Imfeld et al. 2008b, 2010, 2011). Throughout the tertiary and quarternary aquifers, even down to ∼50 m below the surface, chemical and molecular biological markers supported the presence of organohalide respiration and the corresponding bacteria (Imfeld et al. 2011). Analysing a subset of samples, a homogeneous Dehalococcoides sp. affiliated with the Pinellas subgroup was observed in groundwater samples as well as laboratory systems, including microcosms and wetland derived from the groundwater from the edge of the DCE plume at ‘Bergmannshof’ (Mészáros et al. 2013). Interestingly, tceA, coding for the TCE reductive dehalogenase, could not be detected in any of the samples, even though D. mccartyi strain BTF08 was enriched from samples collected from the same site. Strain BTF08 is capable of dechlorinating TCE and possesses the tceA gene (Mészáros et al. 2013; Pöritz et al. 2013). Variable bvcA sequences were detected only in enriched systems such as microcosms and constructed wetlands, while identical vcrA sequences were detected in all types of samples (Mészáros et al. 2013). Furthermore, 12 additional groundwater wells from distant locations within the Bitterfeld megasite exhibiting different patterns of contamination by chlorinated ethenes and benzenes were examined for the presence of known OHRB. Nine wells contained D. mccartyi strains, some of which also contained Dehalobacter and/or Desulfitobacterium based on 16S rRNA gene sequences and all contained at least one of five tested rdhA genes, suggesting active reductive dechlorination was occurring. The three remaining wells were characterised by pH values below 5.5 or a redox potential > +330 mV, pointing to the limits of the natural attenuation at the Bitterfeld site (unpublished data). Chlorobenzene degradation in groundwater Chlorobenzenes, mainly monochlorobenzene, are another relevant contaminant in the Bitterfeld-Wolfen region (Heidrich et al. 2004). Monochlorobenzene may be present as a result of contamination but also could result from dehalogenation of higher chlorinated benzenes or hexachlorocyclohexanes (Field and Sierra-Alvarez 2008; Lal et al. 2010). Dehalococcoides mccartyi strain BTF08, which was enriched from the Bitterfeld groundwater (see below), was observed to reduce γ-HCH to MCB and benzene (Kaufhold et al. 2013), suggesting that this strain could contribute to MCB production in situ. One main anoxic MCB plume was investigated in more detail (see Fig. 1), and in this case a decrease in concentration was associated with an enrichment of 13C in MCB, indicating anaerobic biodegradation (Kaschl et al. 2005). Further studies involving [13C6]-labelled MCB confirmed the utilisation of MCB as a carbon source by indigenous microorganisms under anoxic conditions in both in situ microcosms and laboratory microcosms (Nijenhuis et al. 2007b). The microorganisms and their corresponding pathways have not been resolved yet. Iron oxides, however, were observed to stimulate mineralisation of MCB (Schmidt et al. 2014). In separate experiments, applying [13C6]-MCB as a tracer, DNA-stable isotope probing in combination with the sequencing of bands from single-strand conformation polymorphism analysis allowed to investigate the MCB-metabolising community. This was observed to include bacteria from the phyla Proteobacteria, Fibrobacteres as well as from the candidate division OD1 (Martinez-Lavanchy et al. 2011). To date, further enrichment attempts have failed. Dioxin-dechlorination potential in contaminated freshwater sediments Sediments of different creeks and rivers in the area East and North of Bitterfeld, which are subject to annual high floods, are highly polluted with PCDD/Fs. Total concentrations of 3400 and 7300 ng (kg d.w.)−1 in the Leine-Durchstich creek and the river Mulde (unpublished data), respectively, and 8560 000 ng (kg d.w.)−1 for the Spittelwasser creek (Bunge et al. 2007) were detected. The potential of the intrinsic microbial community to dechlorinate PCDDs was demonstrated with microcosms spiked with 1,2,3,4-tetrachlorodibenzo-p-dioxin (TeCDD) as a model congener (Bunge and Lechner 2001; Bunge, Ballerstedt and Lechner 2001). Lower chlorinated products were detected in microcosms from all sources, including the sulfate-rich acidic (pH 4) Leine-Durchstich sediment (Fig. 1) and six samples from Spittelwasser (representing two neighbouring sampling sites at three different depths). The final dechlorination products were 1,3- and 2,3-dichlorodibenzo-p-dioxin in all cases. Three dechlorination pathways, characterised by different combinations of chlorine removal from lateral and peripheral positions (Fig. 2), were identified in subcultures supplemented with the possible intermediates 1,2,3- and 1,2,4-trichlorodibenzo-p-dioxin (TrCDD), suggesting the presence of diverse dioxin-dechlorinating communities (Bunge and Lechner 2001; Bunge, Ballerstedt and Lechner 2001). Dehalococcoides mccartyi and Desulfitobacterium spp. were detected in these subcultures. In analogy to the dioxin-dechlorinating D. mccartyi strain CBDB1, which was shown to dechlorinate selected dioxin congeners (Bunge et al. 2003), a role of the sediment-derived D. mccartyi population in PCDD dechlorination was suggested and further supported by its growth during dechlorination of 1,2,3- and 1,2,4-TrCDD (Ewald et al. 2007). Figure 2. View largeDownload slide Pathways for the reductive dechlorination of the model congener 1,2,3,4-TeCDD. Sediment microcosms from the Mulde river and the creeks Spittelwasser and Leine-Durchstich dechlorinated 1,2,3,4-TeCDD by different sequences of chlorine removal from peripheral (peri) and lateral positions (solid arrows) resulting in three different patterns of products (see e.g. Adrian and Lechner 2004). Dehalococcoides mccartyi strain DCMB5 isolated from Spittelwasser showed a pathway restricted to the removal of peripheral chlorines only (empty arrows), suggesting that the sediments harbour a much larger diversity of dioxin-dechlorinating bacteria. TrCDD, DCDD, MCDD: tri- di- and monochlorodibenzo-p-dioxin. Figure 2. View largeDownload slide Pathways for the reductive dechlorination of the model congener 1,2,3,4-TeCDD. Sediment microcosms from the Mulde river and the creeks Spittelwasser and Leine-Durchstich dechlorinated 1,2,3,4-TeCDD by different sequences of chlorine removal from peripheral (peri) and lateral positions (solid arrows) resulting in three different patterns of products (see e.g. Adrian and Lechner 2004). Dehalococcoides mccartyi strain DCMB5 isolated from Spittelwasser showed a pathway restricted to the removal of peripheral chlorines only (empty arrows), suggesting that the sediments harbour a much larger diversity of dioxin-dechlorinating bacteria. TrCDD, DCDD, MCDD: tri- di- and monochlorodibenzo-p-dioxin. ENRICHMENT OF D. MCCARTYI STRAINS FROM THE BITTERFELD SITE Dehalococcoides mccartyi strain BTF08: a dedicated chloroethene-respiring groundwater isolate Isolation of microorganisms relevant for the observed activity at a field site is a challenge and associated with many artefacts because laboratory conditions may favour growth of minor populations. Interestingly, a homogeneous D. mccartyi population based on 16S rRNA gene sequence was observed in Bitterfeld groundwater as well as enrichment cultures (Mészáros et al. 2013). Initially, enrichments from groundwater of well BVV3051 contaminated mainly with DCEs were prepared and shown to be capable of the complete reductive dechlorination of tetrachloroethene to ethene with lactate as electron donor (Nijenhuis et al. 2007a). Subsequent transfers, initially with lactate as electron donor and carbon source, later with hydrogen and acetate as electron donor and carbon source, respectively, resulted in a culture dominated by a Dehalococcoides sp. (Cichocka et al. 2010). The novel strain, D. mccartyi strain BTF08, can couple all dechlorination steps of PCE to ethene to energy conservation (Cichocka et al. 2010) and, accordingly, its genome encodes all three necessary enzymes (PceA, TceA and VcrA) (Pöritz et al. 2013). In addition to the chlorinated ethenes, 1,2-dichloroethane, higher chlorinated benzenes and γ-hexachlorocyclohexane were also reductively dechlorinated by this strain (Fig. 3) (Kaufhold et al. 2013). Overall, strain BTF08 appears to be more specialised toward chlorinated alkanes and alkenes and only dechlorinates highly chlorinated benzenes to trichlorobenzenes (Cichocka et al. 2010; Kaufhold et al. 2013; Schmidt, Lege and Nijenhuis 2014). Figure 3. View largeDownload slide Pathways of the reductive dechlorination of chlorinated ethenes and 1,2-dichloroethane (DCA) by D. mccartyi strain BTF08 (solid arrows) and strain DCMB5 (open arrow). The dashed arrow indicates the minor amounts of VC produced during reductive dechlorination of 1,2-DCA by strain BTF08. Figure 3. View largeDownload slide Pathways of the reductive dechlorination of chlorinated ethenes and 1,2-dichloroethane (DCA) by D. mccartyi strain BTF08 (solid arrows) and strain DCMB5 (open arrow). The dashed arrow indicates the minor amounts of VC produced during reductive dechlorination of 1,2-DCA by strain BTF08. Dehalococcoides mccartyi strain DCMB5: a dioxin-dechlorinating isolate from Spittelwasser sediment In an attempt to isolate a dioxin-dechlorinating bacterium from Spittelwasser sediment, 1,2,3-trichlorobenzene (1,2,3-TrCB) was added as an alternative electron acceptor to cultures previously sub-cultured on 1,2,4-TrCDD. 1,2,3-TrCB was applied in a hexadecane phase at high concentrations. It was rapidly dechlorinated to 1,3-dichlorobenzene. A strong increase of bacteria capable to dechlorinate 1,2,4-TrCDD was observed (Bunge et al. 2008). Successive inhibition of methanogens and gram-positive bacteria increased the number of Dehalococcoides cells and finally allowed the isolation of strain DCMB5 from agar shake dilutions (Bunge et al. 2008). This strain showed a strong preference for the dechlorination of peripheral chlorine substituents (Fig. 2) from dioxin congeners and even removed the single chlorine from 1-monochlorodibenzo-p-dioxin (Pöritz et al. 2015), demonstrating its potential contribution to the complete dechlorination of dioxin congener mixtures. Moreover, it was capable of dechlorinating a broad range of chlorinated benzenes and phenols as well as PCE, which it dechlorinated to TCE (Fig. 3). The preference for the dechlorination of chlorinated aromatics might be the result of the enrichment procedure, but it also reflects its origin from heavily dioxin- and chlorobenzene-polluted sediment. Accordingly, its genome does not encode any classical dehalogenase responsible for the dechlorination of chlorinated ethenes such as PceA, TceA or VcrA. Rather, it encodes an ortholog of CbrA, the chlorobenzene reductive dehalogenase in strain CBDB1 (Adrian et al. 2007; Pöritz et al. 2013). As expected, this ortholog, Dcmb_86, was the most abundant dehalogenase in proteomes of pentachlorobenzene- and 1,2,3-TrCB-grown cells, but also of PCE-grown cells, suggesting a broader substrate specificity (Pöritz et al. 2015). Another property of strain DCMB5 is the capability to form type IV pili, detected by transmission electron microscopy of exponentially growing cells. The corresponding gene cluster is conserved in D. mccartyi suggesting its significance for its natural lifestyle. CONCLUSIONS Despite the large horizontal and vertical dimensions of an extremely complex multisource environmental pollution, first insights into an ongoing natural attenuation remediation strategy and related processes at the Bitterfeld megasite were obtained. Based on a relatively dense monitoring network of sampling wells, thorough chemical and hydrogeological analyses (Wycisk et al. 2013), and the selection of site-specific contamination profiles, biomarkers, microbiological and stable isotope analyses were applied and throughout revealed the presence of actively OHRB (see also Table 1). It is interesting to note that D. mccartyi was present at sites contaminated with chloroethenes despite the high concentrations of co-contaminating benzenes and BTEX (Heidrich, Weiß, Kaschl 2004; Nijenhuis et al. 2007a; Cichocka et al. 2010). This suggests that D. mccartyi is relatively resistant to the presence of hydrophobic, non-halogenated contaminants, whereas it is more dependent on suitable ranges of pH and redox potential. This indicates that the natural attenuation of chlorinated ethenes is ongoing at several places within the Bitterfeld megasite and was particularly shown for the DCE plume near the Bergmannshof area. Strain BTF08 isolated from well BVV3051 can be used as an example to simulate the influence of groundwater parameters on the dechlorination process. Its properties reflect well the ecophysiological adaptation of D. mccartyi to this site. The presence of the vcrA gene enabled efficient dechlorination of the main contaminants DCE and VC. It was also able to dechlorinate PCE and TCE in accord with the presence of the respective pceA and tceA genes. The tceA gene was not detected in situ and, thus, strain BTF08 might represent a minor population at the field site (Mészáros et al. 2013). Even though, PCE and TCE were dechlorinated throughout the field site (Imfeld et al. 2011). Overall, D. mccartyi appears to be more capable to adapt to a variety of environmental conditions compared to e.g. Dehalobacter, Geobacter or Desulfuromonas spp., as D. mccartyi was more frequently detected in groundwater samples. For monochlorobenzene, which is one of the organic pollutants with the highest groundwater concentrations in Bitterfeld (>50 000 μg/l at the former chlorobenzene synthesis plant, Wycisk 2003; Kaschl et al. 2005), continuous natural attenuation processes are indicated and have been verified; however, the organisms responsible are not yet known. Although the dioxin-dechlorinating potential was demonstrated for the heavily contaminated Spittelwasser sediment and the isolate D. mccartyi DCMB5, it must be assumed that the natural attenuation process might be slow due to the low water solubility and strong sorptive behaviour of PCCD/Fs. The in situ assessment of PCDD/F dechlorination, e.g. by carbon isotope fractionation analyses, is a difficult task due to the complex congener patterns and the difficulty in extracting sufficient amounts of the compounds. The lower chlorinated products might, however, be a suitable target for analysis (Ewald et al. 2007; Liu et al. 2010). The hexachlorocyclohexane isomers are another problematic contaminant group at the Bitterfeld site, where large deposits represent secondary sources for continuing contamination of groundwater and surface water. Although initial data showed reductive dechlorination of γ-HCH by D. mccartyi strain BTF08 (Kaufhold et al. 2013), and isotopic and enantiomer patterns suggest in situ biodegradation of HCHs in groundwater at the Bitterfeld site (Liu et al. 2017), there is still a gap in our understanding of the intrinsic potential of the microbial communities in groundwater and river sediments and flood plains for the attenuation of HCH isomers, which will need to be addressed in the future. To further address the activity and provide an insight into the ecophysiology of anaerobic microbial dehalogenating bacteria, a suite of methods is already available (Table 1). Currently, however, mainly a direct indicator of activity, such as could be achieved by direct detection of the reductive dehalogenases, via proteomic approaches, is lacking. A combination of methods, such as the analysis of geochemical conditions, substrate and metabolite analysis, compound-specific stable isotope analysis in combination with specific detection of dehalogenating bacteria and their functional genes and a statistical framework, however, may provide such initial insights (Atashgahi et al. 2017). Acknowledgements We would like to thank Marlén Pöritz for sharing unpublished data and Gary Sawers for English proofreading. FUNDING This study was funded by the Deutsche Forschungsgemeinschaft (Research Unit FOR 1530, NI 1329/1-1 and/1-2 as well as LE 780/4-1 and/4-2) and the Helmholtz Centre for Environmental Research—UFZ. Conflict of interest. None declared. REFERENCES Adrian L, Lechner U. Reductive dechlorination of polychlorinated benzenes and dioxins. 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