This is the submitted version of the article: Blázquez Palli, Natàlia; Shouakar-Stash, Ofan; Palau, Jordi; [et al.]. «Use of dual element isotope analysis and microcosm studies to determine the origin and potential anaerobic biodegradation of dichloromethane in two multi-contaminated aquifers». Science of the total environment, Vol. 696 (Dec. 2019), art. 134066. DOI 10.1016/j.scitotenv.2019.134066 This version is avaible at https://ddd.uab.cat/record/212956 under the terms of the license 1 Use of dual element isotope analysis and microcosm 2 studies to determine the origin and potential anaerobic 3 biodegradation of dichloromethane in two multi- 4 contaminated aquifers 5 Natàlia Blázquez-Pallía,b, Orfan Shouakar-Stashc,d,e, Jordi Palauf,g, Alba Trueba- 6 Santisoa, Joan Variasb, Marçal Boschb, Albert Solerf, Teresa Vicenta, Ernest Marco- 7 Urreaa,* and Mònica Rosellf 8 a Departament d’Enginyeria Química, Biològica i Ambiental, Universitat Autònoma de 9 Barcelona (UAB), c/ de les Sitges s/n, 08193 Cerdanyola del Vallès, Spain. b 10 Litoclean, S.L., c/ Numància 36, 08029 Barcelona, Spain. c 11 Department of Earth and Environmental Sciences, University of Waterloo, Waterloo, 12 Ontario N2L 3G1, Canada d 13 Isotope Tracer Technologies Inc., Waterloo, Ontario N2 V 1Z5, Canada e 14 School of Engineering, University of Guelph, Guelph, Ontario, N1G 2W2, Canada f 15 Grup MAiMA, SGR Mineralogia Aplicada, Geoquímica i Geomicrobiologia, 16 Departament de Mineralogia, Petrologia i Geologia Aplicada, Facultat de Ciències de la 17 Terra, Institut de Recerca de l’Aigua (IdRA), Universitat de Barcelona (UB), c/ Martí 18 Franquès s/n, 08028 Barcelona, Spain. g 19 Institute of Environmental Assessment and Water Research (IDAEA), CSIC, Jordi 20 Girona 18-26, 08034 Barcelona, Spain 21 1 * 22 Corresponding author: 23 Departament d’Enginyeria Química, Biològica i Ambiental, Universitat Autònoma de 24 Barcelona (UAB), c/ de les Sitges s/n, 08193 Cerdanyola del Vallès, Spain. E-mail: 25 ernest.marco@uab.cat, Phone: +34 935812694. 26 Abstract 27 Many aquifers around the world are impacted by toxic chlorinated methanes derived from 28 industrial processes due to accidental spills. Frequently, these contaminants co-occur with 29 chlorinated ethenes and/or chlorinated benzenes in groundwater, forming complex 30 mixtures that become very difficult to remediate. In this study, a multi-method approach 31 was used to provide lines of evidence of natural attenuation processes and potential 32 setbacks in the implementation of bioremediation strategies in multi-contaminated 33 aquifers. First, this study determined i) the carbon and chlorine isotopic compositions 34 (δ13C, δ37Cl) of several commercial pure phase chlorinated compounds, and ii) the 35 chlorine isotopic fractionation (εCl = -5.2 ± 0.6‰) and the dual C–Cl isotope correlation 36 (ΛC/Cl = 5.9 ± 0.3) during dichloromethane (DCM) degradation by a Dehalobacterium- 37 containing culture. Such data provide valuable information for practitioners to support 38 the interpretation of stable isotope analyses derived from polluted sites. Second, the 39 bioremediation potential of two industrial sites contaminated with a mixture of organic 40 pollutants (mainly DCM, chloroform (CF), trichloroethene (TCE), and mono- 41 chlorobenzene (MCB)) was evaluated. Hydrochemistry, dual (C–Cl) isotope analyses, 42 laboratory microcosms, and microbiological data were used to investigate the origin, fate 43 and biodegradation potential of chlorinated methanes. At Site 1, δ13C and δ37Cl 44 compositions from field samples were consistent with laboratory microcosms, which 45 showed complete degradation of CF, DCM and TCE, while MCB remained. 46 Identification of Dehalobacter sp. in CF-enriched microcosms further supported the 2 47 biodegradation capability of the aquifer to remediate chlorinated methanes. At Site 2, 48 hydrochemistry and δ13C and δ37Cl compositions from field samples suggested little 49 DCM, CF and TCE transformation; however, laboratory microcosms evidenced that their 50 degradation was severely inhibited, probably by co-contamination. A dual C-Cl isotopic 51 assessment using results from this study and reference values from the literature allowed 52 to determine the extent of degradation and elucidated the origin of chlorinated methanes. 53 54 55 56 3 57 1. Introduction 58 Dichloromethane (DCM) is a probable human carcinogen (IARC, 2016), included in the 59 list of priority pollutants of both the U.S. Agency for Toxic Substances and Disease 60 Registry (2016) and the European Commission (2012). DCM can be naturally released 61 by oceanic sources, wetlands, volcanoes and macroalgae (Gribble, 2010), but its detection 62 in groundwater is often a result of its extensive use, accidental spills and improper storage 63 in chemical, pharmaceutical, and petroleum industrial facilities, among others (Marshall 64 and Pottenger, 2016). 65 Numerous bacteria capable of using DCM as a growth substrate under aerobic 66 conditions have been identified, including various strains of Hyphomicrobium (Heraty et 67 al., 1999; Hermon et al., 2018; Nikolausz et al., 2006) and Methylobacterium 68 (Torgonskaya et al., 2019). The aerobic pathway for DCM biodegradation begins with a 69 glutathione S-transferase producing formaldehyde, which is partially oxidized to CO2 and 70 chloride (Muller et al., 2011). To date, anaerobic biodegradation of DCM has been solely 71 reported for bacteria affiliated with the Peptococcaceae family: Dehalobacterium, 72 Dehalobacter, and Candidatus Dichloromethanomonas elyunquensis (D. elyunquensis). 73 Dehalobacterium formicoaceticum (Dhb f.) is the only pure culture described using DCM 74 as carbon and energy source under anoxic conditions, and produces formate and acetate 75 as end products (Mägli et al., 1998). Recently, mixed bacterial cultures containing 76 Dehalobacter sp. and/or D. elyunquensis have been reported to ferment DCM to mainly 77 acetate (Chen et al., 2017; Justicia-Leon et al., 2012; Kleindienst et al., 2017; Lee et al., 78 2015, 2012). 79 Besides direct release into groundwater, DCM can also be produced by 80 dehalogenation of the higher chlorinated methanes (CMs), i.e. carbon tetrachloride (CT) 81 or chloroform (CF). The latter is listed as a probable carcinogen and a priority 4 82 contaminant as well (ATSDR, 2016; European Commission, 2012; IARC, 2018). Under 83 anoxic conditions, DCM might derive either from the hydrogenolysis of CF by the 84 organohalide-respiring bacteria Dehalobacter sp. and Desulfitobacterium sp. (Chan et al., 85 2012; Ding et al., 2014; Tang and Edwards, 2013; Wong et al., 2016). Also, the anaerobic 86 degradation of CF by Acetobacterium sp. comprises both a reductive branch leading to 87 DCM and an oxidative pathway leading to CO2 (Egli et al., 1990, 1988; Wanner et al., 88 2018). DCM can be also produced co-metabolically by, for example, Methanosarcina sp. 89 or Clostridium sp. (Egli et al., 1988; Gälli and McCarty, 1989; Krone et al., 1989b, 1989a; 90 Mikesell and Boyd, 1990). 91 Chlorinated solvents are commonly found in complex mixtures in groundwater from 92 contaminated sites rather than as individual chemicals. In multi-contaminated aquifers, 93 microbial degradation can be affected and even inhibited by co-contaminants. This 94 information is relevant to foresee the efficiency of natural attenuation or enhanced 95 bioremediation strategies. For instance, the inhibitory effects of CF on key microbial 96 processes such as methanogenesis and microbial dechlorination reactions of chlorinated 97 ethenes (CEs) and ethanes are well known (Weathers and Parkin, 2000; Wei et al., 2016). 98 In addition, it has also been reported that CT and CF inhibit their mutual biodegradation 99 (da Lima and Sleep, 2010; Grostern et al., 2010; Justicia-Leon et al., 2014). 100 Compound-specific isotope analysis (CSIA) is a useful tool that allows 101 bioremediation practitioners to assess the effectiveness of remediation treatments. During 102 chemical or biological degradation of target contaminants, bonds containing the lighter 103 isotopes are preferentially broken, causing the remaining contaminant to be enriched in 104 the heavier isotopes compared to the original isotopic value (Aelion et al., 2009). This is 105 quantified through the abundance ratio of specific stable isotopes (e.g. 13C/12C, 37Cl/35Cl) 106 in contaminant molecules relative to an international standard (Aelion et al., 2009). This 5 107 technique may be used for source apportionment as well as the monitoring of 108 transformation processes in the field (Elsner, 2010). In addition, a quantitative estimation 109 of contaminant transformation extent in the field can be possible, provided that the isotope 110 fractionation (ε) for a given compound and degradation pathway are known and well 111 constrained (Aelion et al., 2009). 112 Dual element isotope analysis has some advantages over a single element isotope 113 approach. On the one hand, by improving the identification of the source (either related 114 to commercial solvents release or from parent compound degradation); on the other, by 115 allowing the elucidation of the fate of pollutants in the field. During biodegradation 116 processes, single element intrinsic isotope effects related to C–Cl bond cleavage can be 117 masked due to the occurrence of previous rate-limiting steps such as preceding enzymatic 118 reactions (Sherwood Lollar et al., 2010), bioavailability of electron donor and/or acceptor 119 (Aeppli et al., 2009; Kampara et al., 2008; Thullner et al., 2008), substrate uptake and 120 transport through the cell membrane (Cichocka et al., 2007; Renpenning et al., 2015), 121 among others. These processes are generally non- or slightly-isotope fractionating so that 122 both elements (e.g. C and Cl) are affected similarly. In this case, by taking the ratio of the 123 isotope shift for the two elements (e.g. Λ = 13C / 37Cl), their masking effect is 124 cancelled out and these slopes reflect better the ongoing degradation mechanisms (Elsner, 125 2010). While single element isotope fractionation could provide insight into the 126 underlying reaction mechanisms in laboratory biodegradation experiments, this is not 127 possible under field conditions. The reason is that contaminant concentration changes at 128 the field are also related to processes other than its transformation (such as sorption and 129 hydrodynamic dispersion), preventing accurate calculation of ε values. Thus, Λ values 130 determined from laboratory studied reactions can be compared with those obtained from 131 groundwater samples to investigate degradation processes at the field scale (Badin et al., 6 132 2014; Hermon et al., 2018; Hunkeler et al., 2009; Palau et al., 2017; Rodríguez-Fernández 133 et al., 2018b). 134 Several laboratory studies have applied dual C–Cl isotope analysis to describe biotic 135 and abiotic transformation mechanisms for DCM (Chen et al., 2018; Heraty et al., 1999; 136 Torgonskaya et al., 2019) and CF (Heckel et al., 2019, 2017a; Rodríguez-Fernández et 137 al., 2018a; Torrentó et al., 2017) but, to the best of the author’s knowledge, the application 138 of a dual isotope approach in DCM-contaminated sites has not been reported yet. 139 The purpose of the work presented herein is two-fold. First, it aims to determine i) the 140 ΛC/Cl value for the anaerobic degradation of DCM by a Dehalobacterium-containing 141 culture, and ii) the carbon and chlorine isotopic compositions (δ13C, δ37Cl) of several 142 commercial pure phase DCM and CF. These new isotope data enrich the database to 143 which dual C–Cl isotopes slopes measured in the field can be compared to elucidate 144 degradation mechanisms. Second, a multi-method approach was used at two industrial 145 sites impacted by DCM, CF, trichloroethene (TCE) and mono-chlorobenzene (MCB) to 146 investigate their origin, fate and intrinsic biodegradation potential. To this end, 147 hydrochemical conditions in groundwater, concentrations of target contaminants and their 148 stable isotope ratios (δ13C, δ37Cl) in field samples were analysed. Afterwards, a dual C– 149 Cl isotopic assessment using the results from this study and reference values from the 150 literature was performed to reveal the origin and fate of DCM. Lastly, field-derived 151 microcosms testing for biostimulation as well as bioaugmentation, and 16S rRNA high- 152 throughput sequencing were performed to assess the intrinsic biodegradation potential of 153 target contaminants at the sites. 154 155 2. Materials and methods 7 156 2.1. Materials 157 Pure phase chlorinated compounds isotopically characterised and/or used for the 158 experiments were purchased at the highest purity available from the following suppliers: 159 CF (Sigma Aldrich, Alfa, Merck), DCM (Sigma Aldrich, Fisher, SDS, Merck), MCB 160 (Sigma Aldrich) and TCE (Panreac). CF (Sigma Aldrich) was used for the enrichment of 161 field-derived cultures. Industrial grade sodium L-lactate (97% purity, at 60% w/w) was 162 from Purac (Corbion). All other chemicals and reagents used were of the highest purity 163 available. 164 Samples of the time-course degradation of DCM in a culture containing 165 Dehalobacterium (Dhb) were obtained from a previous study reported elsewhere 166 (Trueba-Santiso et al., 2017). 167 Two different commercial bacterial consortia were used for the microcosm 168 bioaugmentation tests. Detailed information about the products cannot be disclosed due 169 to commercial confidentiality reasons. Commercial inoculum “S” was reported to 170 degrade CF to DCM, and fully dechlorinated CEs to ethene (ETH). Commercial inoculum 171 “M” was reported to degrade CF to methane via DCM and chloromethane. 172 2.2. Hydrogeochemistry of the studied sites 173 The sites are located in the Catalan Coastal Ranges, which are characterized by an 174 echelon fault system, subparallel to the coast, that affected the Hercynian basement and 175 the Mesozoic overlying sediments. The Neogene extension created a "horst and graben" 176 system filled by Miocene and Quaternary sediments. Both sites are located in the Neogene 177 basin, one of them in the Miocene sediments and the other in the granitic basement. 178 Site 1 8 179 The studied site is located at an industrial area in the Barcelona province (Spain). 180 Geologically, it is constituted by clays and conglomerates of quaternary age overlying 181 unconformable upper Miocene siliciclastic sediments that dip 15–20º in the NW 182 direction. Petrologically, the Miocene sediments are represented by clay and sands 183 composed of abundant quartz grains, mica and kaolinized feldspars and, locally, by some 184 conglomerates. Hydrogeologically, Miocene aquifer has an average transmissivity of 5– 185 180 m2/d, less than the alluvial aquifer. A geological cross section of the studied area can 186 be found in Section A of the Supporting Information (SI). 187 A previous characterization of the site detected high loads of contaminants and dense 188 nonaqueous phase liquid (DNAPL) in groundwater, which was originated after improper 189 storage of the substances (Figure 1A, SI section A). The main organic halogenated 190 compounds included DCM, CF, TCE, and MCB, but cis-1,2-dichloroethene (cis-DCE), 191 vinyl chloride (VC), tetrachloroethene (PCE), acetone, BTEX, and tetrahydrofuran were 192 also detected. The contamination plume was considered finite and contained. After the 193 initial site characterization, a pump and treat (P&T) remediation system (groundwater 194 extraction) was implemented. Under this ongoing treatment, groundwater flowed radially 195 towards the extraction points and a total of seven conventional fully screened monitoring 196 wells were sampled (Figure 1A, SI section B). 197 Site 2 198 This site is also located at an industrial zone in the Barcelona province (Spain), but 199 more than 30 km apart from Site 1. Geologically, it is constituted by Quaternary alluvial 200 sediments represented by sands and gravels that overlie unconformably the basement of 201 the Neogen basin made by later-Hercynian intrusive granites. Hydrogeologically, 202 groundwater flows mainly from the alluvial gravels to the river freshwater and, to a lesser 203 extent, through the more fractured and weathered granite. A conceptual site model of the 9 204 studied area can be found in SI section C. 205 Like in Site 1, it was previously confirmed that the aquifer in Site 2, originally 206 contaminated because of improper management, contained DNAPL (Figure 2A, SI 207 section C). The most abundant contaminants in groundwater included DCM, CF, TCE, 208 and MCB, but acetone, toluene, PCE, CT, VC, tetrahydrofuran and benzene were also 209 detected in groundwater. After the initial site characterization, dual-phase extraction 210 (DPE) and P&T remediation systems were implemented and ongoing during this study. 211 For the hydrochemical characterization, nine conventional fully screened monitoring 212 wells were sampled (Figure 2A, SI section D). At that time, the groundwater flow 213 direction was E–W, and radially towards the extraction points due to DPE and P&T, and 214 the contamination source was reported to be still active (SI section D). 215 2.3. Collection of groundwater samples 216 Hydrochemical parameters (Eh, pH, T and electric conductivity) were measured in- 217 situ and groundwater samples from selected monitoring wells (see SI sections B and D) 218 were collected as described elsewhere (Blázquez-Pallí et al., 2019). It has to be taken into 219 account that these values are an average of the screen length. Samples for chemical and 220 isotopic analysis were collected in different sampling containers on July 11th and 12th, 221 2017, from Site 1 and Site 2. Groundwater was collected with a peristaltic pump or bailer, 222 depending on the water table depth. Samples for C–Cl CSIA were preserved with HNO3 223 (pH~2) (Badin et al., 2016) to prevent biodegradation processes. For the microcosm 224 experiments, groundwater with fine sediments from PZ-9, PZ-19 and PZ-36 from Site 2 225 (SI section D) and PI-2 from Site 1 (SI section B) was sampled on the 18th and 19th of 226 June 2018, respectively. All samples were collected in amber glass bottles sealed with 227 PTFE caps and stored in the dark at 4ºC until used. 10 228 2.4. Set up of laboratory microcosms 229 The microcosms were prepared within the following two days after sampling. They 230 consisted of 65 mL of sampled groundwater in 100 mL glass serum sterile bottles sealed 231 with Teflon-coated butyl rubber septa as described elsewhere (Blázquez-Pallí et al., 232 2019). To investigate whether a potential bioremediation treatment would be feasible at 233 the investigated sites, three different experiments were prepared in triplicate: (i) a control 234 containing only groundwater from the site, which tested for monitored natural attenuation 235 (MNA); (ii) groundwater with sodium lactate (~4 mM), which tested for biostimulation; 236 and (iii) groundwater inoculated with the commercial bacterial consortia described above 237 (106–107 cells/mL) plus sodium lactate (~4 mM), which tested for bioaugmentation. For 238 the samples from wells PZ-19 and PZ-36 (Site 2, SI section D), two different 239 bioaugmentation tests were established, i.e. one with each commercial inoculum, whereas 240 only one bioaugmentation test (using the commercial inoculum “S”) was performed with 241 samples from PZ-9 (Site 2, SI section D) and PI-2 (Site 1, SI section B). 242 To identify the bacteria responsible for CF degradation, the dilution-to-extinction 243 method (Löffler et al., 2005) was applied in 12-mL vials containing an anoxic defined 244 medium described elsewhere (Martín-González et al., 2015) and using CF as electron 245 acceptor. After four extinction series, the more diluted vial showing activity against CF 246 was used as inoculum for serum bottle microcosms, which were selected for 16S rRNA 247 analysis after consuming 1 mM CF. The serum bottle microcosms were prepared as 248 described elsewhere (Martín-González et al., 2015). 249 2.5. DNA extraction and 16S rRNA gene amplicon sequencing 250 Biomass was harvested from whole serum bottle microcosms in sterile falcon tubes 251 that were centrifuged for 40 min at 7000 × g and 10 ºC in an Avanti J-20 centrifuge. The 11 252 pellets were resuspended in sterile PBS buffer and, afterwards, DNA was extracted with 253 the Gentra Puregene Yeast/Bact kit (Qiagen) following the instructions of the 254 manufacturer. Extraction and analysis of the DNA of each microcosm was performed 255 separately. The regions V3–V4 of the 16S rRNA genes were amplified with primers S- 256 D-Bact-0341-b-S-17 and S-D-Bact-0785-a-A-21 (Klindworth et al., 2013) with the 257 Illumina MiSeq platform at Serveis de Genòmica i Bioinformàtica from Universitat 258 Autònoma de Barcelona. 259 2.6. Analytical methods 260 The hydrogeochemical parameters of groundwater (temperature, pH, redox potential 261 (Eh) and electric conductivity) were determined in-situ using a multiparameter probe 262 3430 WTW (Weilheim) as described elsewhere (Blázquez-Pallí et al., 2019). 263 Chlorinated compounds were quantified by analysing 500 µL of headspace samples 264 by gas chromatography (GC) coupled to a flame ionization detector (FID), as described 265 elsewhere (Martín-González et al., 2015). Lactate, acetate and other short-chain fatty 266 acids (VFAs) were monitored by HPLC from 1 mL filtered liquid samples (0.22 μm, 267 Millex), as previously reported (Mortan et al., 2017). 268 Stable carbon isotope ratios (13C) of pure in-house standards were determined with a 269 Flash EA1112 (Carlo-Erba, Milano, Italy) elemental analyser (EA) coupled to a Delta C 270 isotope ratio mass spectrometer (IRMS) through a Conflo III interface (Thermo Finnigan, 271 Bremen, Germany), while the 13C values of target chlorinated compounds (DCM, CF, 272 CEs and MCB) in field samples were obtained by CSIA using an Agilent 6890 GC 273 coupled to a Delta Plus IRMS, as described in Blázquez-Pallí et al. (2019). All analyses 274 (standards and experimental samples) had a one standard deviation (1σ) lower than 0.5‰. 275 Stable chlorine isotope ratios (37Cl) were obtained by CSIA at Isotope Tracer 12 276 Technologies Inc., Waterloo (ON, Canada). A 6890 GC (Agilent, Santa Clara, CA, U.S.) 277 coupled to a MAT 253 IRMS (Thermo Finnigan, Bremen, Germany) were used. This 278 IRMS, equipped with nine collectors, is a continuous flow IRMS with a dual-inlet (DI) 279 mode option. The DI bellows are used as the monitoring gas reservoir (either CF or DCM) 280 and reference peaks were introduced at the beginning of each analysis run (Shouakar- 281 Stash et al., 2006). For the analysis of chlorine isotope ratios of CF, the two main ion 282 peaks (m/z 83 and 85) were used, which correspond to isotopologue pairs that differ by 283 one heavy chlorine isotope ([35Cl212C1H]+ and [37Cl35Cl12C1H]+, respectively) (Breider 284 and Hunkeler, 2014). For DCM, the chlorine isotopic composition was determined from 285 two ion peaks of the molecular group (m/z 84 and 86) corresponding to [35Cl212C1H2]+ 286 and [37Cl35Cl12C1H2]+, respectively. Similarly to carbon isotopes analysis, the analytes 287 were extracted by headspace solid-phase microextraction, as shown elsewhere (Palau et 288 al., 2017). Samples and standards were diluted at similar concentrations and measured in 289 duplicate. The δ37Cl values of pure in-house standards were characterized relative to 290 SMOC (Standard Mean Ocean Chloride) by offline IRMS analysis after conversion of 291 DCM to methyl chloride according to Holt et al. (1997). These standards were later used 292 to correct all measurements from samples. Precision (1σ) of the analysis was ≤ 0.2‰ for 293 37Cl. 294 The isotopic compositions of carbon and chlorine are reported in delta notation (δhE, 295 in ‰), relative to the international standards VPDB (Vienna Pee Dee Belemnite) and 296 SMOC (Coplen, 1996; Kaufmann et al., 1984), respectively (Eq. 1), 𝑅𝑠𝑎𝑚𝑝𝑙𝑒 297 𝛿ℎ𝐸 = ( − 1) (1) 𝑅𝑠𝑡𝑑 298 where Rsample and Rstd are the isotope ratios (i.e. 13C/12C, 37Cl/35Cl) of the sample and the 299 standard of an element E, respectively (Elsner, 2010). 13 300 2.7. Calculations for interpretation of isotopic data 301 A simplified version of the Rayleigh equation in the logarithmic form (Eq. 2) correlates 302 changes in the isotopic composition of an element in a compound (Rt/R0) with changes in 303 its concentration for a given reaction (f = Ct/C0), which allows to determine the 304 corresponding isotopic fractionation (ε) (Coplen, 2011; Elsner, 2010). 𝑅𝑡 305 ln ( ) = 𝜀 · ln(𝑓) (2) 𝑅0 306 Rt/R0 can be expressed as (δhEt + 1) / (δhE0 + 1) according to δhE definition. For the 307 experiment of DCM degradation by the stable enrichment culture containing Dhb, the εCl 308 was obtained from the slope of the linear regression according to Eq. 2. The uncertainty 309 corresponds to the 95% confidence interval (CI). 310 For a target contaminant, the extent of degradation (D%) in the field was evaluated 311 with Eq. 3, which is derived from the Rayleigh equation (Eq. 2), 1000 𝛿 ℎ 𝐸𝑡 + 1000 𝜀𝐸 312 𝐷% = [1 − ( ℎ ) ] · 100 (3) 𝛿 𝐸0 + 1000 313 where εE refers to the isotopic fractionation, δhEt are the isotope data from groundwater 314 samples and δhE0 is the most depleted value found at the field site, assumed to be the most 315 similar to the original source. To this regard, differences in isotope values in the field for 316 both carbon and chlorine (Δδ13C, Δδ37Cl) must be >2‰ for degradation to be considered 317 significant (Hunkeler et al., 2008). Laboratory derived ε values can either be site-specific 318 (i.e., from microcosm experiments prepared with soil and/or groundwater from the 319 contaminated site) or obtained from the literature (Table 1). 320 Finally, the ΛC/Cl value for the DCM biodegradation experiment was obtained from 321 the slope of the linear regression in the dual C–Cl isotope plot (Elsner, 2010) and the 14 322 uncertainty is the 95% CI. 323 2.8. Dual C–Cl isotopic assessment based on selected ΛC/Cl 324 To understand the origin and fate of DCM and CF at each site, a dual C–Cl isotopic 325 assessment (Badin et al., 2016; Hunkeler et al., 2008) was performed based on the 326 measured δ13C and δ37Cl data of field and commercial compounds, the definition of D% 327 (Eq. 3), and εC, εCl and ΛC/Cl from the available literature (Tables 1 and 2). The procedure 328 followed for the dual C–Cl isotopic assessment is detailed in SI section E. 329 330 3. Results and discussion 331 3.1. δ13C and δ37Cl of pure in-house commercial standards 332 Several pure in-house commercial standards were analysed to determine their δ13C and 333 δ37Cl compositions (see details in Table 2). δ13CDCM for the standards that were measured 334 in this study ranged from -35.4 ± 0.2 to -41.71 ± 0.08‰, while δ37ClDCM ranged from +2.1 335 ± 0.3 to -2.7 ± 0.1‰. δ13CCF ranged from -47.96 ± 0.06 to -52.7 ± 0.1‰, while δ37ClCF 336 ranged from -2.6 ± 0.1 to -5.9 ± 0.2‰. A significant variation was observed among the 337 different DCM and CF commercial standards analysed in this study, but also when 338 compared to some of the values available in the literature. For instance, the δ37ClDCM 339 value obtained for the Fisher standard (-2.7 ± 0.1‰) is more depleted in 37 Cl than the 340 values available to date (Table 2), emphasizing the relevance of reference isotopic 341 composition of the pure products provided in this study for future data interpretation. 342 Lastly, δ13C for TCE and MCB was also analysed here and resulted in -25.40 ± 0.03‰ 343 and -27.09 ± 0.07‰, respectively. 344 3.2. Chlorine isotope fractionation and ΛC/Cl during anaerobic DCM degradation by a 345 Dhb-containing culture 15 346 Samples of a Dhb-containing culture that were previously killed to analyse the 347 decrease of concentration and the carbon isotope fractionation of DCM (Trueba-Santiso 348 et al., 2017) were used in this study. The δ37Cl during DCM degradation increased from 349 -5.7 ± 0.2 to +7.2 ± 0.2‰ after 97% transformation, but neither changes in concentration 350 nor in δ37Cl (average -5.7 ± 0.2‰) were observed in the abiotic controls. However, this 351 last point (97%) was excluded because it appeared to be deviated from the linear 352 regression (R2=0.95, Eq. 2), whereas the best-fitting was obtained for δ37Cl values up to 353 84% degradation (R2=0.98) leading to a εCl of -5.2 ± 0.6‰ (Figure 3A). This 354 phenomenon was observed previously by Mundle et al. (2013), who reported higher 355 uncertainty in ε values calculated at later stages of the reaction and suggested examining 356 the linearity of the fits over shorter reaction progress intervals. To date, there is still no 357 εCl value available for Dehalobacter, but the εCl value of -5.2 ± 0.6‰ obtained in this 358 study for Dhb is not significantly different from those reported for D. elyunquensis (-5.2 359 ± 0.1‰), Dhb f. (-5.3 ± 0.1‰), and purified dehalogenases of M. extorquens DM4 and 360 Methylophilus leisingeri DM11(-5.7 and -5.6‰, respectively) (Chen et al., 2018; 361 Torgonskaya et al., 2019) according to the student’s t-test (8 degrees of freedom, p=0.05). 362 In contrast, it is significantly different than those reported for Hyphomicrobium sp. strain 363 MC8b (-3.8‰) and Methylobacterium extorquens DM4 (-7.0‰) (Heraty et al., 1999; 364 Torgonskaya et al., 2019) (Table 1). 365 The εC recalculated up to 84% degradation for Dhb (-31 ± 3‰, R2=0.986) does not 366 vary much from the previously published value (-27 ± 2‰, R2=0.985) (Trueba-Santiso et 367 al., 2017), and still differ from the reported DCM-fermentative bacteria Dehalobacter (- 368 16 ± 2‰) (Lee et al., 2015). However, recent studies on Hyphomicrobium sp. strains 369 (Hermon et al., 2018) and Methylobacterium extorquens DM4 (Torgonskaya et al., 2019) 370 has widened the εC range attributed to methylotrophic bacteria (-22 to -66.6‰), not 16 371 allowing to distinguish hydrolytic transformation of DCM via glutathione-dependent 372 dehalogenases and fermentation pathway just by carbon isotopes. 373 A very good linear correlation (R2=0.995) was also obtained for δ13C against δ37Cl 374 values up to 84% degradation in a dual C–Cl isotope plot (ΛC/Cl of 5.9 ± 0.3, Figure 3B). 375 This ΛC/Cl value by the Dhb-containing culture is significantly different from those 376 reported for D. elyunquensis (ΛC/Cl = 3.40 ± 0.03) and for Dhb f. (ΛC/Cl = 7.89 ± 0.12) 377 (Chen et al., 2018) (Figure 3B) according to the student’s t test (8 degrees of freedom, 378 p=0.05). The difference between Dhb and D. elyunquensis is in accordance with recent 379 proteogenomic findings suggesting that both genera have distinct DCM degradation 380 pathways (Kleindienst et al., 2019). However, the difference observed between Dhb and 381 Dhb f. may suggest that they either have mechanistic disparities during DCM degradation, 382 even though they both belong to the same genus. Higher ΛC/Cl values, ranging from 8.1 383 to 11.2, have been reported for the aerobic Hyphomicrobium sp. strain MC8b (Heraty et 384 al., 1999), Methylobacterium extorquens DM4, and DM4 and DM11 DCM dehalogenases 385 (Torgonskaya et al., 2019). To date, these values hint at the distinction between aerobic 386 (11.2 to 8.1) and anaerobic (7.89 to 3.40) pathways (Figure 3B). However, further 387 research on this topic would be needed to confirm such hypothesis. See Table 1 for a 388 detailed review of literature values. 389 390 3.3. Field sites investigation 391 392 3.3.1. Site 1 393 The studied aquifer exhibited mixed redox conditions (Eh ranged from +250 to -52) 394 mV), but most of the wells were in anoxic conditions (SI section F), which are conducive 17 395 to reductive dechlorination reactions. Temperature and pH were considered homogeneous 396 throughout the aquifer (21.1 ± 0.7 ºC and 6.8 ± 0.3, respectively) (SI section F). CF was 397 the organohalide detected at the highest concentration (454 mg/L), followed by MCB 398 (82.7 mg/L), TCE (47.2 mg/L), and DCM (1.47 mg/L) (Figure 1A and SI section G). 399 The C–Cl isotopic composition of CF could be determined in 5 out of 7 monitoring 400 wells (Figure 1A). Obtained δ13C values for CF showed a relatively small but significant 401 variation (Δδ13CCF) of 3.2‰ between wells PI-17 and CV-5, which exhibited the most 402 enriched and depleted δ13CCF, respectively (SI section H). Since the difference in the 403 carbon isotope composition was ˃2‰, it could be related to CF transformation in 404 groundwater (Hunkeler et al., 2008). In contrast, the maximum Δδ37ClCF observed at the 405 site between wells CV-5 and PI-24 was lower than 2‰. This agrees well with previous 406 studies showing εC values usually greater than εCl during CF biodegradation (Table 1). 407 The dual isotopic composition of CF in the analysed groundwater samples fell outside the 408 isotopic range of commercial CF, except for samples collected in well CV-5 (SI section 409 I). This result would further support a potential transformation of CF at the site. 410 For DCM, only one C–Cl isotopic composition could be determined (in well PI-26) 411 due to DCM concentration being too low for carbon CSIA, but δ37Cl could be determined 412 in 4 out of 7 wells (Figure 1A). Obtained δ37Cl for DCM showed a significant variation 413 in Cl/35Cl ratios of Δδ37ClDCM = 2.3‰ between wells CV-5 and PI-26 (SI section H, 37 414 Figure 1A), and these δ37Cl were enriched compared to those of commercial DCM (Table 415 2). As mentioned earlier, no Δδ13CDCM can be provided for δ13C since only one value 416 could be obtained, but this value was also enriched compared to most of the δ13C reported 417 for commercial DCM (Table 2). Hence, this dual C–Cl isotopic composition for DCM in 418 the analysed groundwater of well PI-26 fell outside the isotopic range of commercial 419 DCM, suggesting a potential transformation of DCM at the site as well (SI section I). 18 420 When comparing these results to the literature, the low carbon enrichment observed could 421 indicate that the bacteria responsible for DCM degradation have a εC that is in the lower 422 range (e.g. Dehalobacter sp., which has the lowest reported εC as presented in Table 1). 423 An additional evidence pointing towards in-situ DCM biodegradation was the detection 424 of acetate in wells PI-26 and CV-3 (Figure 1A and SI section G), which is the main by- 425 product of DCM fermentation (Justicia-Leon et al., 2012; Lee et al., 2015; Mägli et al., 426 1998; Trueba-Santiso et al., 2017). 427 Interestingly, δ13CTCE in the wells that also contained CF ranged between -32.5 and - 428 31.8‰ but was significantly enriched in 13 C in the unique well (PI-3) that lacked CF 429 (δ13CTCE = -28.2‰, Figure 1A). This may be in agreement with several studies suggesting 430 that CF can inhibit significantly microbial reductive dechlorination of CEs, as reviewed 431 elsewhere (Wei et al., 2016). However, cis-DCE, VC and ETH were also detected in most 432 of the monitored wells (SI section G) indicating that, despite this potential inhibition 433 effect suggested by the isotopic analysis, full dechlorination of TCE occurred. 434 Lastly, no significant carbon isotope fractionation was observed for MCB (Δδ13CMCB 435 < 2‰, SI section H), and obtained δ13C values were within the available range of 436 commercial MCB solvents (Table 2), indicating that this compound was, most likely, not 437 being degraded. 438 To investigate the biodegradation potential of the site, three different microcosm 439 treatments were prepared with groundwater from well PI-2 (SI section J). The non- 440 amended microcosms used as natural attenuation controls fully degraded CF and DCM, 441 and transformed TCE to VC after 75 days. The microcosms amended with either only 442 lactate (biostimulation) or also inoculated with “S” (bioaugmentation) increased the 443 degradation rate of CF (to DCM), DCM, TCE and, more importantly, fully dechlorinated 444 the latter to stoichiometric amounts of ETH. Both CMs and CEs elimination occurred 19 445 simultaneously (SI section J). These results confirmed the feasibility of in-situ CF, DCM 446 and TCE biodegradation, and agreed with the information obtained from the isotopic field 447 data that was just discussed. In contrast, MCB always accumulated in the medium of the 448 microcosms and was not degraded (SI section J). Such evidence further supports the 449 evidence obtained from the isotopic field data, which pointed towards the recalcitrance 450 of the pollutant. Thus, it could be discarded that benzene, which was detected in the 451 preliminary characterization of this aquifer, was derived from MCB biodegradation. 452 Regarding the amended substrate, lactate was converted to acetate (major product) and 453 propionate in all the amended microcosms (SI section K). 454 The composition of the microbial community was assessed by 16S rRNA high- 455 throughput sequencing. The analysis was performed on groundwater samples collected 456 from well PI-2 that were enriched with CF for four consecutive dilution series. The 457 taxonomic assignments of 16S rRNA sequences revealed that the four predominant 458 genera were Clostridium (31%), Sedimentibacter (25%), Dehalobacter (10%) and 459 Acetobacterium (4%) (SI section L). The presence of Dehalobacter, a well-known CF 460 and DCM degrader (Justicia-Leon et al., 2012; Lee et al., 2015), further supports the 461 intrinsic biodegradation potential of the aquifer to remediate CMs. In addition, results 462 also suggested that Clostridium sp. and Acetobacterium sp. could be somehow involved 463 in the degradation of CF as well, as described elsewhere (Egli et al., 1990, 1988; Gälli 464 and McCarty, 1989; Wanner et al., 2018). 465 With these evidences and aiming to better understand the origin and fate of CF and 466 DCM at this site, a dual C–Cl isotopic assessment was performed based on the assumption 467 that CF in well CV-5 reflected the initial isotopic signature (CF0) since it possessed the 468 most depleted δ13C and δ37Cl values (Figure 1A). The dual isotope fractionation pattern 469 observed for CF in CV-3, PI-17 and PI-26 was consistent with the degradation of leaked 20 470 commercial CF, since the data points fell within the estimated potential isotopic 471 composition for degraded CF (orange area, Figure 1B). Accordingly, the values for DCM 472 initially produced from CF degradation (DCMini) ranged between -50.9‰ and -73.9‰ 473 for δ13C and was of -2.2‰ for δ37Cl (Figure 1B, SI section E). The isotopic composition 474 of DCM detected in well PI-26 (SI section H) was enriched in 37Cl compared to that of 475 commercial DCM (Table 2), and in both 13C and 37Cl compared to that of DCM produced 476 by CF degradation (SI section E), suggesting that it could be subject to biodegradation 477 (Figure 1B). However, regarding the main source of detected DCM at the site, obtained 478 results point to a potential spill of DCM at the site being more likely than DCM produced 479 by CF degradation, but neither could be confirmed. 480 In the above scenario, and assuming Dehalobacter sp. was the responsible for CF 481 biotransformation, the extent of biodegradation (D%) for CF estimated in the wells CV- 482 3, PI-17 and PI-26 (where Δδ13C was > 2‰, SI section H) ranged between 7–11% 483 considering the εC data reported for Dehalobacter strain CF (εC = -28 ± 2)). Data for 484 Dehalobacter strain CF was used for this calculation because its ΛC/Cl was more coherent 485 with the field results from Site 1 (i.e. the isotopic pattern of positive dual C–Cl slopes 486 observed for CF measurements), than the ΛC/Cl of Dehalobacter strain UNSWDHB 487 (Table 1). Nevertheless, further research needs to be done on this culture to better 488 understand the roles of Dehalobacter, Clostridium and Acetobacterium bacteria, and 489 elucidate the biodegradation mechanisms of both CF and DCM, which would allow an 490 improved assessment and quantification of D% at the field. 491 3.3.2. Site 2 492 At this site, Eh ranged from +254 to -118 mV, but groundwater in most of the wells 493 exhibited oxic conditions which limit the occurrence of reductive dechlorination reactions 494 (SI section F). Temperature and pH were considered homogeneous and averaged 24 ± 2 21 495 ºC and 6.8 ± 0.3, respectively (SI section F). CF was detected at the highest concentrations 496 (272 mg/L), followed by DCM (27 mg/L), TCE (1.7 mg/L) and MCB (0.88 mg/L) (Figure 497 2A and SI section G). 498 δ13C and δ37Cl of CF could be determined in 7 out of 9 wells because the concentration 499 of CF in PZ-5 and PZ-31 was below the limit of quantification (Figure 2A). The Δδ13CCF 500 > 2‰ observed between PZ-9 and PZ-18 wells and the Δδ37ClCF < 2‰ observed between 501 PZ-10 and PZ-18 wells (SI section H) could be related to CF transformation in 502 groundwater (Hunkeler et al., 2008) considering the low εCl reported for CF 503 transformation. However, most of the C–Cl isotope compositions of CF were within the 504 isotopic range for commercial CF (SI section I), which could also indicate that measured 505 δ13C and δ37Cl belonged to a leaked and not degraded CF. 506 On its side, detected DCM presented a higher variation for both C and Cl isotope ratios 507 (Δδ13CDCM = 6.1‰ between wells PZ-19 and PZ-36, Δδ37ClDCM = 2.2‰ between wells 508 PZ-9 and PZ-18, Figure 2A, SI section H) indicative of potential in-situ transformations. 509 In this case, however, the isotopic composition in well PZ-36 did fall within the 510 commercial DCM, whereas well PZ-19 did not (Figure 2B and SI section I, see further 511 discussion below). 512 For TCE, Δδ13C was of 5.8‰ (SI section H), which could be related to TCE 513 transformation in groundwater (Hunkeler et al., 2008), and measured δ13CTCE values were 514 more enriched in C than the δ13C compositions of commercial TCE (Table 2), in 13 515 agreement with potential TCE degradation. However, the reductive dechlorination 516 products of TCE (DCE, VC, and ETH) were detected occasionally and at very low 517 concentrations (SI section G). 518 Microcosm experiments were prepared with groundwater from wells PZ-9, PZ-19 and 22 519 PZ-36. Microcosms for well PZ-9 did not show significant degradation of neither CF nor 520 CEs, and no major differences were observed between the non-amended, biostimulated 521 and the bioaugmented treatments after 150 days (SI section M). However, traces of DCM 522 and VC, a slight decrease in PCE and TCE, and formation of trans-DCE were observed 523 throughout the course of the study (SI section M). For PZ-19, no significant elimination 524 of contaminants was observed, although TCE did exhibit a decrease in all microcosms 525 and traces of cis-DCE, trans-DCE and VC were detected in some of them (SI section N). 526 For PZ-36, the non-amended, biostimulated and bioaugmented (with “M”) microcosms 527 did neither degrade CF nor CEs significantly. In contrast, the bioaugmented (with “S”) 528 treatment exhibited a CF decrease with the transient production of DCM, and TCE was 529 completely transformed to VC and trans-DCE (SI section O). Amended lactate was 530 mainly transformed to acetate in PZ-9 and PZ-36 microcosms. In PZ-19, however, lactate 531 was not completely consumed but, in both bioaugmentation tests, its decrease was 532 followed by acetate and propionate production (SI section K). The inability of inoculated 533 bacteria to efficiently degrade contaminants in these microcosms suggested that there 534 were inhibition issues, possibly due to the presence of co-contaminants in groundwater 535 such as CT, which is a known inhibitor of the microbial activity and CF degradation in 536 particular (section 2.2) (da Lima and Sleep, 2010; Grostern et al., 2010; Justicia-Leon et 537 al., 2014; Wei et al., 2016). The results obtained with the microcosms suggested that, 538 under the studied conditions, degradation of contaminants was very slow and inefficient, 539 but possible to some extent, which agreed with the information obtained from the isotopic 540 field data that was discussed above. 541 At this site, the dual C–Cl isotopic assessment was performed based on a CF0 signature 542 of δ13C = -45.1‰ (from PZ-9) and δ37Cl = -3.5‰ (from PZ-10), as they were the most 543 depleted values measured (Figure 2A). The dual isotope fractionation pattern observed 23 544 for CF was not consistent with degradation (orange area, Figure 2B), in agreement with 545 the microcosms. However, it should be noted that δ37ClCF values were enriched respect 546 to CF0 signature. This could be attributable to Cl isotope fractionation processes in the 547 unsaturated zone due to diffusion-controlled vaporization (Jeannottat and Hunkeler, 548 2012; Palau et al., 2016) or to the leakage of CF from different providers with distinct 549 isotopic signatures. Moreover, it has to be taken into account that this release of CF was 550 still active at the site and potential CF degradation could be masked by a more depleted 551 CF input. In the hypothetical case that CF was biologically degraded, the calculated 552 isotopic range for DCMini (formation of DCM from CF degradation) would range from - 553 52.1‰ to -75.1‰ for carbon, and -3.5‰ for chlorine (Figure 2B, SI section E). The 554 isotopic composition of DCM detected in groundwater (Figure 2B, SI section H) was very 37 13 555 enriched in both Cl and C compared to that of initial DCM produced by CF 556 degradation (grey area, Figure 2B). In detail, DCM detected in PZ-36 (with the highest 557 DCM concentration, Figure 2A) was within the range of commercial DCM (Figure 2B), 558 pointing to a potential release of DCM at the site. However, this DCM could be subject 559 to a little biodegradation as acetate, the main by-product of DCM fermentation, was 560 detected in the well PZ-36 and also in PZ-18 (SI section G). For the well PZ-19, the 561 δ13CDCM value is slightly depleted in 13C while the δ37ClDCM value is lightly enriched in 37 562 Cl compared to those measured in PZ-36 (Figure 2B). This could be explained similar 563 to CF, in this case DCM could be an impurity in the CF raw source with isotopic signature 564 changing over time due to change in CF providers or manufacturing processes or by the 565 effect of vaporization and diffusion processes in the unsaturated zone (Jeannottat and 566 Hunkeler, 2012; Palau et al., 2016). As noted above, the release of CF (and probably 567 DCM) was still active at the site. Taking all these evidences into account, the extent of 568 biodegradation (D%) for CF was not estimated since the results would not be 24 569 representative of CF degradation. 570 571 4. Conclusions 572 The present study provides additional data on isotope C and Cl compositions for 573 different commercial CMs, widening the known range for these solvents and proving the 574 utility of these results for data interpretation. In addition, the C and Cl isotopic 575 fractionation values for the anaerobic degradation of DCM by a Dhb-containing culture 576 were also determined. These results enrich the available database that can be used by 577 practitioners to provide diagnostic information about CMs biodegradation in 578 contaminated aquifers. The value of ΛC/Cl obtained for the investigated Dhb-containing 579 culture was right in between of those described for Dhb f. and D. elyunquensis, which 580 would allow the distinction of these DCM degradation mechanisms through isotope 581 analyses. 582 The pumping regimes that were active at both studied sites probably impacted the 583 geographical distribution of the chlorinated compounds and their corresponding isotopic 584 signatures. For this reason, the focus of this work was not on the flow path of the plume 585 but on the correlation between the isotopic composition of the chlorinated solvents and 586 the biodegradation information obtained from microcosms experiments for each 587 monitoring well. For this, the use of an integrative approach that combines C–Cl CSIA, 588 laboratory microcosms, and 16S rRNA high-throughput sequencing demonstrated the 589 intrinsic biodegradation potential of Site 1 to fully transform CF, DCM, and CEs, but not 590 MCB. In contrast, results for Site 2 suggested that inhibition was preventing the efficient 591 elimination of CMs and CEs at tested conditions. Nevertheless, the dual C–Cl isotopic 592 assessment proved useful, as well, to elucidate the origin and fate of DCM and CF at both 25 593 sites. Considering these results, a biostimulation (e.g. enhanced reductive dechlorination 594 with lactate) could be applied at Site 1 to degrade both CMs and CEs. However, the 595 persistence of MCB at this site, and the inhibition of biodegradation observed at Site 2, 596 suggest that a treatment train (i.e. a combination or sequence of different remedial 597 strategies targeting different groups of contaminants) could possibly be the optimal 598 approach for the detoxification of these aquifers. 599 This study shows that such a multi-method approach allows for the collection of data 600 that can help making decisions in the field. In this case, it is a valuable tool to evaluate 601 the feasibility of biodegradation strategies to remediate chlorinated solvents in complex 602 multi-contaminated aquifers, as it can provide the lines of evidence required to 603 demonstrate whether bacteria can successfully detoxify groundwater, and identify any 604 potential setbacks. The microcosm tests can provide a positive indication that complete 605 dechlorination can be achieved by native microbial populations, and is useful to predict 606 the effect of bioremediation treatments (biostimulation or bioaugmentation) to detoxify 607 the aquifer. Changes in carbon and chlorine isotope composition among the residual 608 fraction of the chlorinated compounds in the monitoring wells can provide field evidence 609 for microbial degradation. Lastly, the positive molecular identification of key 610 organohalide-respiring bacteria (e.g. Dehalobacter) can provide additional evidence that 611 chlorinated solvents can be fully dechlorinated in the aquifer. Notwithstanding, there is a 612 need for additional laboratory studies that correlate the metabolism of microbial 613 transformations with stable isotope fractionation of multiple elements to support the 614 interpretation of CSIA data obtained from contaminated groundwaters, as well as the 615 potential inhibitory effect of co-contaminants over bacteria degrading organochlorides. 616 617 5. Acknowledgements 26 618 This research has been supported by the Spanish State Research Agency (CTM2016- 619 75587-C2-1-R and CGL2017-87216-C4-1-R projects), co-financed by the European 620 Union through the European Regional Development Fund (ERDF). This work was also 621 partly supported by Generalitat de Catalunya through the consolidate research groups 622 (2017SGR-14 and 2017SGR-1733) and the Industrial Doctorate grant of N. Blázquez- 623 Pallí (2015-DI-064). A. Trueba-Santiso acknowledges support from a MICINN 624 predoctoral research grant (BES-2014-070817). M. 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Detailed concentrations (in µM) and isotopic compositions are available in 917 SI sections G and H, respectively. “n.d.” means “could not be determined based on their 918 low concentration”; “n.a.” means “not analysed”. (B) Dual C–Cl isotopic assessment for 919 DCM and CF field data from Site 1. Both CF0 and the range for DCMini are represented. 920 Green and orange dashed rectangles depict the δ13C and δ37Cl ranges of commercial DCM 921 and CF solvents, respectively (see Table 2 for details). 922 41 923 Figure 2. (A) Concentrations (in mg/L) and δ13C and δ37Cl compositions (in ‰) of CF, 924 DCM, TCE and MCB from Site 2. Pie charts show the molar distribution of contaminants 925 at each well. Detailed concentrations (in µM) and isotopic compositions are available in 926 SI sections G and H, respectively. “n.d.” means “could not be determined based on their 927 low concentration”; “n.a.” means “not analysed”. (B) Dual C–Cl isotopic assessment for 928 DCM and CF field data from Site 2. Both CF0 and the range for DCMini are represented. 929 Green and orange dashed rectangles depict the δ13C and δ37Cl ranges of commercial DCM 930 and CF solvents, respectively (see Table 2 for details). 931 42 932 Figure 3. Double logarithmic Rayleigh plot for chlorine isotope data (A) and dual C–Cl 933 isotope plot (B) for the anaerobic degradation of DCM by the Dhb containing culture 934 investigated in this study. Solid black lines for the Dhb culture in each panel depict the 935 corresponding linear regression and dashed lines represent the associated 95% CI. In 936 panel B, the trend lines reported for DCM degradation by Hyphomicrobium strain MC8b 937 (Heraty et al., 1999), Dhb f. and D. elyunquensis (Chen et al., 2018) are shown for 938 comparison. Data points show the error bars from duplicate measurements. 939 940 43 941 TABLES 942 Table 1. εC, εCl and ΛC/Cl of DCM and CF degradation by several mechanismsa. DCM εC (‰) εCl (‰) ΛC/Cl Reference This study and re-calculated from Trueba- Mixed culture containing Dehalobacterium sp. (Dhb) -31 ± 3* -5.2 ± 0.6 +5.9 ± 0.3 Santiso et al. (2017)* Dehalobacterium formicoaceticum (Dhb f.) -42.4 ± 0.7 -5.3 ± 0.1 +7.89 ± 0.12 Chen et al. (2018) Consortium RM harboring Ca. Dichloromethanomonas elyunquensis -18.3 ± 0.2 -5.2 ± 0.1 +3.40 ± 0.03 Chen et al. (2018) Mixed culture containing Dehalobacter sp. (DCMD) -16 ± 2 n.a.b n.a. Lee et al. (2015) Hyphomicrobium sp. strain MC8b -42.4c,d -3.8c,d +11.2c,d Heraty et al. (1999) Hyphomicrobium strains, Methylobacterium, -41.2 to -66.3d n.a. n.a. Nikolausz et al. (2006) Methylopila, Methylophilus, and Methylorhabdus strains -45.8 to -61.0e Nikolausz et al. (2006) Hyphomicrobium strains -22 to -46d n.a. n.a. Hermon et al. (2018) -26 ± 5e Hermon et al. (2018) Methylobacterium extorquens DM4 (average between high and low -66.6d,f -7.0d,f +9.5d,f Torgonskaya et al. (2019) density cell suspensions) DM4 DCM dehalogenase (average between high and low activity) -48.4d,f -5.7d,f +8.5d,f Torgonskaya et al. (2019) DM11 DCM dehalogenase (average between high and low activity) -45.4d,f -5.6d,f +8.1d,f Torgonskaya et al. (2019) CF εC (‰) εCl (‰) ΛC/Cl Reference Oxidation -8 ± 1 -0.44 ± 0.06 +17 ± 2 Torrentó et al. (2017) Alkaline hydrolysis -57 ± 5 -4.4 ± 0.4 +13.0 ± 0.8 Torrentó et al. (2017) Hydrogenolysis plus reductive elimination with Fe (0) -33 ± 11 -3 ± 1 +8 ± 2 Torrentó et al. (2017) Biodegradation with vitamin B12 -14 ± 4 -2.4 ± 0.4 +7 ± 1 Rodríguez-Fernández et al. (2018a) Outer-sphere single electron transfer (OS-SET) -17.7 ± 0.8 -2.6 ± 0.2 +6.7 ± 0.4 Heckel et al. (2017a) Dehalobacter strain CF -28 ± 2 -4.2 ± 0.2 +6.6 ± 0.1 Heckel et al. (2019) Dehalobacter strain UNSWDHB -3.1 ± 0.5 +2.5 ± 0.3 -1.2 ± 0.2 Heckel et al. (2019) Mixed culture containing Dehalobacter strain UNSWDHB -4.3 ± 0.5 n.a. n.a. Lee et al. (2015) aUncertainties of ε and Λ values correspond to the 95% confidence interval (CI). bn.a., values were not analysed. cΛC/Cl values were calculated based on reported εC and εCl data by the referenced authors. dValues were measured under oxic conditions. eValues were measured under nitrate-reducing conditions. fεC and εCl values were calculated here based on reported αC and αCl data (εC,Cl = 1/αC,Cl – 1) · 1000), and the ΛC/Cl values from the here estimated εC and εCl values (ΛC/Cl ~ εC/εCl). 943 44 944 Table 2. δ13C and δ37Cl (in ‰) of commercial pure-phase CF, DCM, TCE and MCB. DCM δ13C (‰/VPDB) δ37Cl (‰/SMOC) Reference Fisher -41.71 ± 0.08 -2.7 ± 0.1 This study Sigma Aldrich - +0.75 ± 0.05 This study SDS -35.4 ± 0.2 +1.27 ± 0.01 This study Merck -39.0 ± 0.1 - This study unknown -39.78 ± 0.08 - This study unknown - +1.3 ± 0.2 This study unknown - +2.1 ± 0.3 This study unknown -31.5 ± 0.3 +2.13 ± 0.03 Jendrzejewski et al. (2001) unknown -31.8 ± 0.5 +2.3 ± 0.2 Jendrzejewski et al. (2001) unknown -34.19 ± 0.02 +1.555 ± 0.005 Holt et al. (1997) unknown -40.4 ± 0.5 - Holt et al. (1997) CF δ13C (‰/VPDB) δ37Cl (‰/SMOC) Reference Mercka -48.93 ± 0.08 - This study Mercka -52.7 ± 0.1 - This study Sigma Aldrich - -2.6 ± 0.1 This study Alfa -47.96 ± 0.06 -5.9 ± 0.2 This study Alfa -47.88 ± 0.08 -5.4 ± 0.3 Breider (2013) unknown -43.21 ± 0.04 -1.52 ± 0.01 Holt et al. (1997) Rodríguez-Fernández et al. (2018b), Fluka -48.7 ± 0.1 -3.0 ± 0.2 Heckel et al. (2017b) Acros -49.76 ± 0.08 - Rodríguez-Fernández et al. (2018b) unknown -51.7 ± 0.4 +0.32 ± 0.08 Jendrzejewski et al. (2001) Fisher -53.23 ± 0.09 - Rodríguez-Fernández et al. (2018b) Sigma Aldrich -63.6 ± 0.1 - Breider (2013) TCE δ13C (‰/VPDB) δ37Cl (‰/SMOC) Reference unknown -25.40 ± 0.03 - This study unknown -27.18 ± 0.01 -1.42 ± 0.10 Holt et al. (1997) PPG -27.37 ± 0.09 -2.8 ± 0.1 Shouakar-Stash et al. (2003) unknown -27.90 ± 0.08 +2.0 ± 0.1 Jendrzejewski et al. (2001) StanChem -29.1 ± 0.1 -3.19 ± 0.07 Shouakar-Stash et al. (2003) ICI -31.01 ± 0.09 +2.71 ± 0.08 Shouakar-Stash et al. (2003) Dow -31.57 ± 0.01 +3.55 ± 0.05 Shouakar-Stash et al. (2003) Dow -31.90 ± 0.05 +4.1 ± 0.3 van Warmerdam et al. (1995) Sigma Aldrich -33.49 ± 0.08 +3.8 ± 0.1 Jendrzejewski et al. (2001) MCB δ13C (‰/VPDB) δ37Cl (‰/SMOC) Reference Sigma Aldrich -27.09 ± 0.07 - This study Sigma Aldrich -25.6b - Liang et al. (2011) a Solvent with the same reference number but from different batches. bδ13C (‰) value for MCB was obtained from a microcosm experiment where changes in δ13C by biodegradation of MCB were measured. The initial δ13C is not considered to be affected by degradation and, therefore, can be assumed as representative of the commercial MCB. VPDB: Vienna Pee Dee Belemnite, SMOC: Standard Mean Ocean Chloride. 945 45