This is the author's version of a work that was accepted for publication in Science of the total environment (Ed. Elsevier). Changes resulting from the publishing process, such as peer review, editing, corrections, structural formatting, and other quality control mechanisms may not be reflected in this document. Changes may have been made to this work since it was submitted for publication. A definitive version was subsequently published in Rosell, Mònica et al. “Dual carbon- chlorine isotope fractionation during dichloroelimination od 1,1,2-trichloroethane by an enrichment culture containing Dehalogenimonas sp” in Science of the total environment, Vol 648, (January 2019), p 422-429. The final versión is available at DOI 10.1016/j.scitotenv.2018.08.071 1 Graphical abstract 2 3 4 5 6HIGHLIGHTS (3-5 bullets; 85 characters per bullet): 7  Dehalogenimonas transforms 1,1,2-trichloroethane (1,1,2-TCA) to vinyl 8 chloride. 9  Dual C-Cl isotope analysis applied for the first time for 1,1,2-TCA degradation. 10  Significant C and Cl isotope fractionation during 1,1,2-TCA 11 dichloroelimination. 12  Calculated Λ can allow distinguishing 1,1,2-TCA degradation pathways in the 13 field. 14 1 1 2 15 Dual carbon - chlorine isotope fractionation during 16 dichloroelimination of 1,1,2-trichloroethane by an enrichment 17 culture containing Dehalogenimonas sp. 18 19Mònica Rosell1, Jordi Palau*1,2, Siti Hatijah Mortan3,a, Gloria Caminal4, Albert Soler1, 20Orfan Shouakar-Stash5,6, Ernest Marco-Urrea3 21 221 Grup MAiMA, Mineralogia Aplicada, Geoquímica i Geomicrobiologia, Departament 23de Mineralogia, Petrologia i Geologia Aplicada, Facultat de Ciències de la Terra, 24Universitat de Barcelona (UB), Martí Franquès s/n, 08028 Barcelona, Spain. 252 Institute of Environmental Assessment and Water Research (IDAEA), CSIC, and 26Hydrogeology Group (UPC−CSIC), Jordi Girona 18-26, 08034 Barcelona, Spain 273 Departament d’Enginyeria Química, Biològica i Ambiental, Universitat Autònoma de 28Barcelona (UAB), 08193 Bellaterra, Barcelona, Spain. 294 Institut de Química Avançada de Catalunya (IQAC), CSIC, Jordi Girona 18-26, 08034 30Barcelona, Spain 315 Department of Earth and Environmental Sciences, University of Waterloo, Waterloo, 32Ontario N2L 3G1, Canada. 336 Isotope Tracer Technologies Inc., Waterloo, Ontario N2 V 1Z5, Canada 34a Current address: Faculty of Chemical and Natural Resources Engineering, Universiti 35Malaysia Pahang, Lebuhraya Tun Razak, Gambang, 26300 Kuantan, Pahang, Malaysia. 36 37*Corresponding author: jordi.palau@ub.edu 38 3 2 4 39ABSTRACT 40 41Chlorinated ethanes are frequent groundwater contaminants but compound specific 42isotope analysis (CSIA) has been scarcely applied to investigate their degradation 43pathways. In this study, dual carbon and chlorine isotope fractionation was used to 44investigate for the first time the anoxic biodegradation of 1,1,2-trichloroethane (1,1,2- 45TCA) using a Dehalogenimonas-containing culture. The isotopic fractionation values 46obtained for the biodegradation of 1,1,2-TCA were ɛC = -6.9 ± 0.4‰ and ɛCl = -2.7 ± 470.3‰. The detection of vinyl chloride (VC) as unique byproduct and a closed carbon 48isotopic mass balance corroborated that dichloroelimination was the degradation 49pathway used by this strain. Combining the values of δ 13C and δ37Cl resulted in a dual 50element C-Cl isotope slope of Λ=2.5 ± 0.2‰. Investigation of the apparent kinetic 51isotope effects (AKIEs) expected for cleavage of a C-Cl bond showed an important 52masking of the intrinsic isotope fractionation. Theoretical calculation of Λ suggested 53that dichloroelimination of 1,1,2-TCA was taking place via simultaneous cleavage of 54two C-Cl bonds (concerted reaction mechanism). The isotope data obtained in this study 55can be useful to monitor natural attenuation of 1,1,2-TCA via dichloroelimination and 56provide insights into the source and fate of VC in contaminated groundwaters. 57 58Keywords: Dehalogenimonas; dual isotope fractionation; dichloroelimination; 59organohalide-respiring bacteria; 1,1,2-trichloroethane. 60 61 5 3 6 621. Introduction 631,1,2-Trichloroethane (1,1,2-TCA) has been widely used as a solvent and chemical 64intermediate in the industry (Pankow and Cherry, 1996). Improper storage and 65accidental spills have contributed to 1,1,2-TCA being a frequent detected contaminant 66in groundwater at industrial facilities (ATSDR, 1989). In the United States, it is ranked 67166 out of 275 substances on the Priority List of Hazardous Substances based on a 68combination of its frequency, toxicity, and potential for human exposure (ATSDR, 692015). 70 Quantification of the distribution and fate of chlorinated contaminants and 71degradation products in the subsurface is a complex task since biological, chemical, and 72physical processes may affect them (Němeček et al., 2017). Biological transformation 73of 1,1,2-TCA is influenced by the intrinsic heterogeneity of natural environments that 74allows for different redox conditions to occur either spatially or temporally separated. 75Under anoxic conditions, reductive dechlorination is expected to be the prevailing 76mechanism to transform 1,1,2-TCA by two different biodegradation pathways: 77hydrogenolysis and dichloroelimination. In the case of dichloroelimination, two vicinal 78C-Cl bonds of 1,1,2-TCA are cleaved to produce vinyl chloride (VC), whereas during 79hydrogenolysis 1,1,2-TCA is sequentially transformed by single C-Cl bond cleavage to 801,2-dichloroethane (1,2-DCA) and monochloroethane (Moe et al., 2016; Zhao et al., 812015) (Fig. 1). The key organisms catalyzing hydrogenolysis and dichloroelimination 82are organohalide-respiring bacteria (OHRB), which can use 1,1,2-TCA as respiratory 83electron acceptor (Leys et al., 2013). To date, dichloroelimination of 1,1,2-TCA has 84been described for OHRB belonging to the genus Dehalobacter and Dehalogenimonas 85(Grostern and Edwards, 2006; Mortan et al., 2017; Yan et al., 2009), but 86hydrogenolysis only for Desulfitobacterium (Zhao et al., 2015). Under oxic conditions, 7 4 8 87no bacteria are currently known to use 1,1,2-TCA as growth substrate, but 88cometabolism of 1,1,2-TCA can occur during aerobic oxidation of methane, propane, 89butane, n-pentane, n-hexane or ammonia (Frascari et al., 2006, 2008, 2013; Vannelli et 90al., 1990). The only byproducts identified for aerobic cometabolism of 1,1,2-TCA 91include chloroacetic acid (which was sequentially oxidized to glyoxylic acid) and minor 92amounts of VC in microcosms containing a Pseudomonas sp. (Castro and Belser, 931990). Abiotic transformation of 1,1,2-TCA can produce a wide array of byproducts, 94including VC (Patterson et al., 2016), ethane (Song and Carraway, 2005) or 1,1- 95dichloroethene (1,1-DCE) (Pagana et al., 1998) (Fig.1). 96 Knowledge on degradation pathways occurring in an aquifer contaminated with 971,1,2-TCA is a key aspect to design suitable bioremediation strategies. However, this is 98a challenge when the site contains multiple chlorinated aliphatic hydrocarbons because 99the same daughter products of 1,1,2-TCA dechlorination can be formed from other 100precursors (i.e., VC is produced from anaerobic reductive dechlorination of DCE 101isomers or 1,2-DCA) (Hunkeler et al., 2002). It is important to note that VC, produced 102during biotic or abiotic reductive dichloroelimination of 1,1,2-TCA, is even much more 103toxic than 1,1,2-TCA. 104 Compound-specific isotope analysis (CSIA) has emerged in recent years as a 105technique with great potential to elucidate specific reaction pathways even if no 106products are detected (Elsner, 2010). The magnitude of carbon and chlorine kinetic 107isotope effects (KIEs) during contaminant degradation relies on the observation that 108lighter stable isotopes (i.e., 12C, 35Cl) react at faster rates than the heavier ones (i.e., 13C, 10937Cl). For a given compound and reaction, single element isotope fractionationvalues 110(are determined in laboratory degradation experiments according to the Rayleigh 111equation. However,  values associated to biodegradation cannot be accurately 9 5 10 112measured in the field because other processes such as sorption or mixing through 113dispersion also affect contaminant concentration. 114 Two-dimensional CSIA brings the potential to overcome the limitation of single 115element isotope fractionation values to identify contaminant degradation pathways in 116the field. Combined changes in isotope ratios of two elements (i.e., 13C and 37Cl) 117for a given reactant generally correlate in a dual element isotope plot obtaining a slope 118( = 13C / 37Cl) that reflects the isotope effects of both elements. Hence,  values 119may act as direct indicator for different initial reaction mechanisms. To interpret dual 120element CSIA data sets obtained from contaminated field sites, it is necessary to know 121experimental carbon and chlorine isotope enrichment factors and values derived from 122microbial strains catalyzing known transformation reactions (Cretnik et al., 2013; 123Kuntze et al., 2016). However, to our knowledge, chlorine isotope fractionation (Cl) 124and values are not available for 1,1,2-TCA. Hunkeler et al. (2002) showed that 125dichloroelimination of 1,1,2-TCA to VC in anaerobic microcosms inoculated with 126contaminated groundwater was accompanied of a relatively weak carbon isotopic 127fractionation of 1,1,2-TCA (εC = -2.0 ± 0.2‰). Recently, in a laboratory flow-through 128column experiment consisting of both biodegradable organic carbon and zero valent 129iron, εC changed from -14.6±0.7‰ to -0.72±0.12‰, being this last value assigned to 130anaerobic biodegradation (Patterson et al., 2016). 131 The main aims of this research were to measure for the first time dual C-Cl 132isotope fractionation and to determine the resultant value during biodegradation of 1331,1,2-TCA with an anaerobic bacterial culture containing a Dehalogenimonas sp. This is 134valuable information i) to investigate the fate of 1,1,2-TCA in future biodegradation 135field studies and ii) to get insight into the underlying reaction mechanism involved in 11 6 12 136the dechlorination of 1,1,2-TCA. In addition, carbon isotope values of VC were 137measured to determine the product isotope pattern during biodegradation of 1,1,2-TCA. 138 1392. Materials and methods 1402.1. Biodegradation batch experiments 141A stable enrichment culture containing a Dehalogenimonas sp. described previously that 142transforms 1,1,2-TCA to VC via dichloroelimination (Martín-González et al., 2015) 143was used in batch experiments. Each microcosm consisted of 100 mL glass serum 144bottles containing 65 mL of a sterilized anoxic synthetic medium previously used to 145grow Dehalococcoides mccartyi strain CBDB1 (Adrian et al., 2000). This medium 146contained vitamins, trace elements, Na2S × 9 H2O and L-cysteine (0.2 mM each) as 147reducing agent, and as carbon source either sodium acetate (5 mM) or pyruvate (5 mM) 148as indicated. The serum bottles were sealed with Teflon-coated butyl rubber septa and 149aluminum crimp caps and gassed with N2/CO2 (4:1, v/v, 0.2 bar overpressure) and H2 150(added to an overpressure of 0.4 bar). 1,1,2-TCA was added with a syringe from a stock 151solution in acetone to give an initial aqueous phase concentration of  20 mol L-1; 152higher concentrations appeared to be inhibitory for this Dehalogenimonas-containing 153culture. 154 A total of 16 parallel incubations from the same inoculum were prepared at the 155same time. Half of these cultures contained acetate and the other half pyruvate as carbon 156source. Cultures were incubated at 25°C in the dark without shaking. Samples were 157collected for isotopic and concentration analyses at different extents of 1,1,2-TCA 158dechlorination. In order to control losses, abiotic transformations, and the transfer of 159compounds with the inoculum (previous growth in 1,2-dichloropropane, 1,2-DCP) or 160potential impurities from the stock solution, two types of controls were included in 13 7 14 161triplicate: (i) live controls without 1,1,2-TCA and (ii) abiotic controls containing the 162growth medium with 1,1,2-TCA but without inoculum. 1632.2. Analytical methods 1642.2.1. Concentration and isotopic measurements. 1,1,2-TCA and VC concentrations 165in serum bottles were monitored along the experiment by taking 0.5 mL headspace (HS) 166samples with a 1.0 mL pressure-lock precision analytical syringe (Vici, U.S.) and 167injecting them in a gas chromatograph (GC) model 6890N (Agilent Technologies) 168equipped with a DB-624 column (30 m × 0.32 mm with 0.25 μm film thickness; Agilent 169Technologies) and a flame ionization detector (FID), as described elsewhere (Palau et 170al., 2017). Depending on the measured concentrations (expressed in μmol L –1 of liquid 171volume) the bottles were sacrificed at different extent of degradation stopping biological 172activity by adding 12 mL of an oxic, saturated H2SO4/Na2SO4 solution (pH=1). 173 Compound-specific carbon and chlorine isotope analyses were performed by 174HS-solid-phase micro-extraction (HS-SPME)-GC-isotope ratio mass spectrometry (GC- 175IRMS) as described elsewhere (Palau et al., 2017). δ13C analyses were performed in the 176Centres Científics i Tecnològics de la Universitat de Barcelona (CCiT-UB), Spain, 177while δ37Cl were carried out at Isotope Tracer Technologies Inc. (IT2), Canada. For 178analyzing chlorine isotope ratios of 1,1,2-TCA, the two most abundant fragment ions 179(m/z 97 and 99) were used, which correspond to isotopologue pairs (i.e., [ 35Cl212C21H3]+ 180and [37Cl35Cl12C21H4]+, respectively) that differ by one heavy chlorine isotope. For 1,1,2- 181TCA, the intensities of the most abundant fragment ion peaks are much higher than 182those of the parent ion peaks. The raw δ37Cl values were calibrated to the standard mean 183ocean chloride (SMOC) scale using a two-point linear calibration. The standards were 184dissolved in water and measured similarly to the samples interspersed in the same 15 8 16 185sequence. Duplicate samples and standards were analyzed. The precision (1σ) on the 186analysis of standards was ≤0.5‰ for δ13C and ≤0.2‰ for δ37Cl. 1872.2.2. Isotope data evaluation. Carbon and chlorine isotope ratios of 1,1,2-TCA were 188measured at natural abundance and were expressed using the -notation in per mil (eq. 1891), 190 E l E /¿ ¿ ¿h ¿ ¿ sample ¿ E 191 l E /¿ (1) ¿ ¿h ¿ R¿ ¿ ¿ Esample =h¿ ¿ δ¿ 192 193where R is the isotope ratio of heavy (hE) to light (lE) isotopes of an element “E “(e.g., 19413C/12C and 37 Cl/35Cl). The relationship between isotope fractionation and the extent of 1951,1,2-TCA biodegradation in laboratory experiments was evaluated by a modified form 196of the Rayleigh distillation equation (2) 197 198 ln ( δ hE S +1 000 δ hE S 0 +1000 ) = ε bulk 1000 ∙ ln f 199 (2) 200 17 9 18 201where hES0 is the initial isotopic composition of element “E” in a substrate “S” and 202hES is the isotopic composition at a remaining fraction “f” (f = C S/CS0). The 203compound-average isotope fractionation values ( ε bulk ) were quantified by least 204squares linear regression of eq. 2 without forcing the regression through the origin 205(Scott et al., 2004) and the uncertainty corresponds to the 95% confidence interval 206(C.I.) derived from the standard deviation of the regression slope. The Rayleigh 207equation can also be applied to calculate the isotopic fractionation of chlorine despite 208the higher natural abundance of 37Cl compared to 13C (Elsner and Hunkeler, 2008). 209 To evaluate the product carbon isotope fractionation pattern, the 13C of VC that 210was produced was calculated using eq. 3, where 13CP is the isotopic composition of the 211product “P” (i.e., VC) and ε bulk is the estimated carbon isotopic fractionation of 2121,1,2-TCA (eq. 2) (Cretnik et al., 2014; Hunkeler et al., 2005). 213 ε bulk ∙ f ∙ ln f 214 δ 13C P=δ 13C S 0− (3) 1−f 215 216 For a given substrate, intrinsic KIEs during compound transformation are 217position specific whereas bulk values are calculated from compound-average isotope 218data (eq. 2). Therefore, observable bulk values must be converted into apparent KIEs 219(AKIEs) in order to obtain information about the underlying reaction mechanisms 220(Elsner et al., 2005). For the calculation and interpretation of AKIEs a hypothesis about 221the reaction mechanism, or assumed reaction mechanism, is necessary. The effects of 222non-reacting positions within the molecule, as well as of intramolecular competition, are 223then taken into account using equations 4 and 5, respectively (Elsner et al., 2005), 224 n 225 ε rp ≈ ∙ε x bulk 226 (4) 19 10 20 1 AKIE C ,Cl = 227 ε z ∙( rp )+1 1000 228 (5) 229 230where rp is the isotopic fractionation at the reactive position, “n” is the number of atoms 231of the element considered, “x” is the number of these atoms at reactive sites (i.e., atoms 232that would experience isotope effects in the given reaction) and “z” the number of 233identical reactive sites undergoing intramolecular competition. These equations assume 234the absence of secondary isotope effects. For carbon, secondary isotope effects are 235usually insignificant (Elsner et al., 2005). For dichloroelimination of 1,1,2-TCA to VC, 236if the two C-Cl bonds are broken in sequence (i.e., stepwise dichloroelimination, single 237C-Cl bond cleavage at the first reaction step), assuming that the first bond cleavage is 238the rate determining step, then n = x = z = 2 and n = x = z = 3 for C and Cl, respectively, 239as all C and Cl atoms are in equivalent position and compete for reaction. On the other 240hand, if the two C-Cl bonds are broken simultaneously (i.e., concerted 241dichloroelimination), the average AKIEC and AKIECl for the two reacting positions were 242calculated since there is no intramolecular competition between them, n = x = 2, z = 1 243and n = 3, x = 2, z = 1 for C and Cl, respectively. AKIEs that were calculated assuming 244stepwise or concerted dichloroelimination are referred hereafter as “ AKIE stepwise ” and 245“ AKIE concerted ” and their uncertainty was calculated by error propagation. 246 For a given substrate and reaction, the dual C-Cl isotope slope () obtained from 24713C vs 37Cl isotope plots can be expressed as follows (Elsner, 2010 and references 248herein): 249 250 ΛC−Cl = ∆ δ 13C ε Cbulk ≈ Cl ≈ x n ( ) ∙ ( A ) KIE −1 ∙ 1+( A ) KIE ∙( z −1) C C C C (6) ( ) ( A ) KIE −1 1+ ( A ) KIE ∙( z −1) 37 ∆ δ Cl ε bulk x Cl Cl Cl n Cl 21 11 22 251 2523. Results and discussion 2533.1. Concentration and isotope patterns 2543.1.1. Dechlorination of 1,1,2-TCA by a Dehalogenimonas-containing culture. The 255anaerobic microcosms amended with pyruvate and acetate as carbon source lasted 256approximately 7 and 15 days, respectively, at which point the initial 1,1,2-TCA was 257transformed to VC via dichloroelimination. The concentration of 1,1,2-TCA in the 258abiotic controls (19.0 ± 0.5 mol L-1, ± 1, n = 5) remained at the initial concentration 259along the experiments, which indicates that compound losses through the caps during 260incubation were insignificant. The difference in the lag phase between acetate and 261pyruvate amended microcosms is not probably associated with the carbon source but to 262the inoculum source that was more enriched in the microcosms with pyruvate. No other 263volatile organic compounds were detected, especially 1,2-DCA was absent discarding 2641,1,2-TCA hydrogenolysis. At different stages of 1,1,2-TCA degradation, isotope 265signatures of 1,1,2-TCA (δ13C and δ37Cl) and VC (δ13C) were measured for all the 266samples to determine the corresponding isotopic fractionation values of 1,1,2-TCA (ε C 267and εCl) and the carbon isotope pattern of produced VC. 2683.1.2. Carbon isotope pattern of 1,1,2-TCA. The δ13C of 1,1,2-TCA in the abiotic 269controls remained constant through both experiments, with a total average value of 270-36.3 ± 0.6‰. In contrast, carbon isotopic composition of 1,1,2-TCA in the cultures 271became progressively enriched in 13C during its degradation reaching a δ13C value up to 272-14.3‰ when 96% of 1,1,2-TCA was degraded in both acetate- and pyruvate-containing 273media (Fig. 2). These results show that despite the differences in the lag phase and the 274inoculum source, no statistical difference in concentrations and carbon isotope values 275was observed for the experiments prepared with either acetate or pyruvate as carbon 23 12 24 276source. Isotopic data from both experiments were combined and the total carbon isotope 277composition of 1,1,2-TCA followed a Rayleigh trend (r 2=0.9901, Fig. 3A) with an εC 278value of -6.9 ± 0.4‰ (95% C.I., n=16). 279 The similar isotope fractionation of 1,1,2-TCA for the microcosms amended 280with either acetate or pyruvate agrees with recent studies investigating isotopic 281fractionation of trichloroethene (TCE) under different growth conditions. Harding et al. 282(2013) showed that carbon isotope fractionation during TCE degradation by 283Dehalococcoides-containing cultures remained consistent despite a variety of 284temperature, nutrient, and cofactor-limiting conditions investigated. In addition, 285Buchner et al. (2015) studied the potential effects of metabolic adaptation on carbon 286and chlorine isotope fractionation of TCE during biodegradation by Desulfitobacterium 287hafniesne Y51. These authors reported similar bulk values for C and Cl isotopes under 288different growth conditions (i.e., cultures pre-grown with fumarate or TCE as electron 289acceptors) and enzyme quantity per cell and suggested that isotope fractionation was not 290affected. 2913.1.3. Carbon isotope pattern of VC. In parallel to 1,1,2-TCA transformation, the δ13C 292of its degradation product (i.e., VC) was monitored. The δ13C of VC was initially 13 293depleted in C, in agreement with the normal isotope effect of 1,1,2-TCA, and shifted 294toward more positive values during the course of reaction reaching the initial value of 2951,1,2-TCA once this was completely degraded (Fig. 2). As observed for 1,1,2-TCA, the 296carbon isotope data of VC from the experiments with acetate and pyruvate showed 297similar values (Fig. 2). This figure also shows that δ 13C values of VC fitted very well 298with the expected product isotope trend determined according to eq. 3. The closed 299isotopic mass balance confirmed the absence of other relevant degradation products. 300Moreover, δ13C of VC never overpass the initial δ13C of 1,1,2-TCA suggesting that VC 25 13 26 301is not further degraded to non-chlorinated compounds such as ethene or ethane, which is 302consistent to its accumulation. 303 A different product isotope pattern was observed for degradation of 1,1,2-TCA 304in a previous study with microcosms constructed with aquifer material and groundwater 305(Hunkeler et al., 2002). These authors observed δ13C values of VC very enriched in 13C 306compared to those of 1,1,2-TCA towards the end of reaction, which was indicative of 307further degradation of VC to ethene via reductive dechlorination. Therefore, the results 308of the present study and Hunkeler et al., 2002 illustrate the potential of the product 309carbon isotope pattern to investigate the fate of VC in sites impacted with 1,1,2-TCA. 310Analysis of ethene concentration can be used to evaluate the fate of VC in groundwater, 311provided that other potential precursors of ethene such as 1,2-DCA are not present at the 312site. However, assessing the fate of VC based solely on ethene concentration can be 313difficult because ethene can be transformed under both oxic and anoxic conditions to 314carbon dioxide and ethane, respectively (Mundle et al., 2012), highlighting the benefit 315of VC isotope analysis as complementary data. 3163.1.4. Chlorine isotope pattern of 1,1,2-TCA and dual C-Cl isotope approach. Chlorine 317isotope data of 1,1,2-TCA (δ37Cl) were obtained from the pyruvate amended 318microcosms. The δ37Cl of 1,1,2-TCA in the abiotic controls (-0.88 ± 0.2‰) did not 319change significantly during the experiment, while an enrichment in the heavy isotope 320(37Cl) during 1,1,2-TCA degradation following a Rayleigh trend (ε Cl = -2.7 ± 0.3‰, n=8, 32195% C.I., Fig. 3B) was observed in the cultures. Chlorine isotope fractionation was 322much lower than for carbon, in agreement with the large primary carbon isotope effects 323expected for C-Cl bond cleavage (Elsner et al., 2005). The measurement of chlorine 324isotope ratios enabled for the first time a dual C-Cl isotope approach for biodegradation 325of 1,1,2-TCA. A very good linear correlation (r2 = 0.994) was obtained when δ13C and 27 14 28 326δ37Cl were combined in a dual element isotope plot showing a slope (Λ) of 2.5 ± 0.2 327(95% C.I., Fig. 4). 328 A recent study on 1,2-DCA showed different Λ values during 329dichloroelimination by Dehalogenimonas- and Dehalococcoides-containing cultures, 330suggesting that a dual C-Cl isotope approach could help to identify the microbial taxa 331responsible for anaerobic biodegradation of 1,2-DCA in the field (Palau et al., 2017). 332This information is particularly important for 1,1,2-TCA given that, in contrast to 333Desulfitobacterium (Zhao et al., 2015) (Fig. 1), its degradation by Dehalogenimonas 334can result in an accumulation of the highly toxic VC in groundwater. Therefore, 335comparison of the Λ value obtained for Dehalogenimonas in the present study with 336those obtained for 1,1,2-TCA degradation by other bacteria in future studies might help 337to investigate the fate of 1,1,2-TCA and to predict potential accumulation of VC in 338contaminated sites. 3393.2. Isotope effects and insight into dichloroelimination mechanisms of 1,1,2-TCA 340Significant variation on reported bulk carbon isotope fractionation during 341biodegradation of 1,1,2-TCA is observed (Table 1). The ε C value of -6.9 ± 0.4‰ 342determined in this study is significantly larger than that previously reported, -2.0 ± 3430.2‰ from microcosms constructed with anaerobic aquifer material and groundwater 344(Hunkeler et al., 2002). In addition, a much lower εC value of -0.7 ± 0.1‰ was 345determined by Patterson et al. 2016, which was attributed to biodegradation in a 346laboratory column consisted of both zero valent iron Fe(0) and biodegradable organic 347carbon. Interestingly, microbiological data from this laboratory column suggested that a 348co-culture composed by Desulfitobacterium and Dehalococcoides was responsible for 349the sequential degradation of 1,1,2-TCA to ethene. The enzymatic mechanism of 350Desulfitobacterium sp. strain PR to transform 1,1,2-TCA to 1,2-DCA via 29 15 30 351hydrogenolysis differs from the production of VC via dichloroelimination in our 352Dehalogenimonas-containing culture which could explain the difference on carbon 353isotope fractionation observed in both studies (Fig. 1). A simultaneous cleavage of two 354C-Cl bonds via concerted dichloroelimination of 1,1,2-TCA might result theoretically in 355a larger bulk εC value compared to hydrogenolysis, where a single C-Cl bond is broken 356at the initial reaction step. However, the occurrence of isotope-masking leading to 357smaller εC values cannot be excluded. In this case, if preceding (rate-limiting) steps 358exhibit small or no isotope fractionation, the observable isotope effect will be smaller 359(i.e., masked) than the intrinsic isotope effect. 360 To address in more detail whether dichloroelimination of 1,1,2-TCA by 361Dehalogenimonas proceeds via a stepwise or concerted mode, AKIE values were 362calculated according to eq. 4 and 5 as it was previously done with the same 363Dehalogenimonas containing enrichment for 1,2-DCP (Martín-González et al., 2015) 364or 1,2-DCA (Palau et al., 2017, see also Table 1). Assuming stepwise or concerted 365mode, carbon AKIEs obtained for 1,1,2-TCA (AKIE Cstepwise = 1.0138 ± 0.0008 and 366AKIECconcerted = 1.0069 ± 0.0004, respectively) were much below the Streitweiser limit of 367KIEC for complete C-Cl bond cleavage (1.057) and the realistic value of 50% bond 368cleavage (1.029) (Elsner et al., 2005), making both modes feasible, but showing 369important masking of intrinsic isotope fractionation. For chlorine, AKIEs determined 370for both mechanisms (AKIEClstepwise = 1.0082 ± 0.0009 and AKIE Clconcerted = 1.0041 ± 3710.0005), were also below the Streitweiser limit for C-Cl bond cleavage (1.013). 372 Apart from theoretical Streitweiser limits, isotopic fractionation values and 373derived AKIEs from abiotic reactions are often considered closest to the intrinsic 374isotope effects. Abiotic reductive dechlorination of 1,1,2-TCA was suggested in the 375same above-mentioned Fe(0) column study but without the organic carbon amendment 31 16 32 376(Patterson et al. 2016). In that case, an AKIE for stepwise mode of 1.0246 can be 377calculated from the reported εC value (−12 ± 5‰). This AKIE value is within the range 378(AKIECstepwise = 1.0158 to 1.0326) previously available for abiotic reductive 379dechlorination of 1,1,1-TCA and other polychlorinated ethanes, 1,1,2,2- 380tetrachloroethane (1,1,2,2-TeCA), pentachloroethane (PCA) and hexachloroethane 381(HCA) by Cr(II), Fe(0) and Cu and Fe mixtures (Elsner et al., 2007; Hofstetter et al., 3822007; Palau et al., 2014). Chlorine isotope effects (AKIEClstepwise = 1.0125 to 1.0207) 383were also reported by Hofstetter et al., 2007 and Palau et al., 2014. The reported 384carbon and chlorine AKIEs for abiotic reductive dechlorination of chlorinated ethanes 385(via single C-Cl bond cleavage at the first reaction step) are higher than those 386determined for 1,1,2-TCA dichloroelimination in this study assuming either stepwise or 387concerted scenarios. Therefore, mechanistic interpretations are challenged by the 388relatively low observable bulk isotope effects of 1,1,2-TCA. The occurrence of isotope- 389masking effects can sometimes complicate the identification of the underlaying reaction 390mechanism since derived AKIEs may then be no longer characteristic of a certain 391reaction (Elsner et al., 2005). However, an improved interpretation might be possible 392by comparing dual C-Cl isotope slopes (see below). 393 A large isotope fractionation masking such that of 1,1,2-TCA during degradation 394by Dehalogenimonas in this study was also observed for Dehalobacter-containing 395mixed culture degrading 1,1,1-TCA versus 1,1-DCA (Sherwood Lollar et al., 2010). In 396particular, the large intrinsic kinetic isotope effect expected for cleavage of a C-Cl bond 397was almost completely masked during 1,1,1-TCA biodegradation by both whole cells 398and cell-free extracts, while for 1,1-DCA the reduction was only roughly 50%. These 399effects were not attributable to transport effects across the cell membrane, rather than to 33 17 34 400significant differences in the kinetics of the enzymes catalyzing chlorinated ethane 401degradation. 4023.3. Reaction mechanism insight from dual C-Cl plot. 403An important advantage of values compared to bulk values (and derived AKIEs) is 404that the magnitude of the latter can be significantly affected by isotope-masking 405processes. Since isotope-masking affect both elements to a similar extent, the dual 406element isotope slopes remain largely unaltered (Elsner, 2010). For 1,1,2-TCA, the lack 407of degradation studies including both carbon and chlorine isotope data makes not 408possible a comparison of the value determined for Dehalogenimonas in this study 409with values for different reactions (biotic and abiotic) and microbial strains. However, 410values for a new compound like 1,1,2-TCA can be predicted based on the expected 411KIEs for carbon and chlorine according to eq 6, and it can be then compared to the 412experimentally determined value for Dehalogenimonas. 413 Assuming concerted dichloroelimination of 1,1,2-TCA (n = x = 2, z = 1 and n = 4143, x = 2, z = 1 for C and Cl, respectively, see above), the carbon and chlorine isotope 415effects determined in a recent study (Palau et al., 2017) for reductive 416dichloroelimination of 1,2-DCA by Dehalogenimonas were used in eq. 6 (AKIE Cconcerted 417= 1.024 ± 0.003 and AKIE Clconcerted = 1.0121 ± 0.0008, see Table 1). These authors 418postulated a concerted character of the reaction based on determined carbon isotope 419effects. As a result, a value of 2.98 was obtained, which is similar to the experimental 420value of 2.5 ± 0.2 (Fig. 4). In contrast, if a stepwise dichloroelimination of 1,1,2-TCA is 421assumed (n = x = z = 2 and n = x = z = 3 for C and Cl, respectively), a very different 422value of 1.01 is obtained. In this case, the average carbon and chlorine isotope effects 423for 1,1,2,2-TeCA, PCA and HCA during abiotic dichloroelimination by Cr(II) via 35 18 36 424sequential -elimination of two chlorine atoms were considered (AKIECstepwise = 1.026 ± 4250.005 and AKIEClstepwise = 1.017 ± 0.004, see Table 1) (Hofstetter et al., 2007). In 426addition, a smaller value of 0.66 was obtained in case the AKIEs estimated for 427reduction of 1,1,1-TCA by Fe(0) via single electron transfer are used in the calculations 428(AKIECstepwise = 1.0158 ± 0.0008 and AKIE Clstepwise = 1.0160 ± 0.0006, see Table 1). 429Therefore, the comparison of the experimental value of 1,1,2-TCA with those 430expected for stepwise and concerted mechanisms according to eq. 6 suggests that a 431concerted dichloroelimination is more likely, highlighting the benefit of using a dual 432C−Cl isotope approach. This result is in agreement with previous studies of 1,2-DCP 433and 1,2-DCA biodegradation by Dehalogenimonas suggesting a concerted 434dichloroelimination pathway (Martín-González et al., 2015; Palau et al., 2017). 435Identification of the underlying transformation mechanism controlling isotope 436fractionation can be valuable information to improve the characterization of reductive 437dehalogenases. In addition, an eventual identification of different dichloroelimination 438mechanisms of 1,1,2-TCA (i.e., concerted vs stepwise) by distinct microbial strains 439might indicate the existence of diverse reductive dehalogenases with similar function 440but likely different structure. For 1,2-DCA, the isotopic differences observed by Palau 441et al. (2017) between Dehalogenimonas and Dehalococcoides containing cultures on 442the concerted dichloroelimination mechanism were associated to a distinct interaction 443mode between cobalamin dependent enzymes rather than two different reaction 444pathways (i.e., stepwise vs concerted). The same isotopic results and conclusions were 445validated by Franke et al. (2017) with two pure Dehalococcoides mccartyi strains (195 446and BTF08). 447 4484. Conclusions 37 19 38 4491,1,2-TCA is a frequent groundwater contaminant but surprisingly only few studies 450applying CSIA have been reported so far. Our work provides the first application of dual 451isotope fractionation to investigate the anaerobic biodegradation of 1,1,2-TCA. The 452stable isotope data obtained in this study during the dichloroelimination of 1,1,2-TCA 453can be potentially helpful in monitoring the fate of this pollutant in contaminated 454environments. In addition, the carbon isotope pattern of VC obtained in our enrichment 455enlightens its potential use to identify the dominant VC production mechanism and 456predict further transformation of this toxic compound. The single element kinetic 457isotope effects could not provide conclusive information about the reaction mechanism 458involved in 1,1,2-TCA dichloroelimination (concerted or stepwise); however, the dual- 459element approach can reduce interpretation bias due to isotope-masking effects 460overcoming this limitation and pointing to more likely concerted mechanism. Further 461investigations on carbon and chlorine isotope fractionation with bacteria catalyzing 462alternate degradation pathways (i.e., hydrogenolysis) will allow the comparison between 463microbial dechlorination reactions of 1,1,2-TCA. 464 4655. Acknowledgements 466This research has been supported by the Spanish Ministry of Economy and 467Competitiveness, MINECO (CTM2013-48545-C2-1-R, CTM2016-75587-C2-1-R, 468CGL2014-57215-C4-1-R and CGL2017-87216-C4-1-R projects) co-financed by the 469European Union through the European Regional Development Fund (ERDF). This work 470was also partly supported by the Generalitat de Catalunya through the consolidate 471research groups (2017-SGR-14 and 2017SGR-1733). M. Rosell acknowledges a Ramón 472y Cajal contract (RYC-2012-11920) from MINECO and S.H. Mortan support from the 473Ministry of Education Malaysia (SLAI-UMP Scholarship) for a predoctoral fellowship. 39 20 40 474The Departament d’Enginyeria Química, Biològica i Ambiental of the Universitat 475Autònoma de Barcelona is a member of the Xarxa de Referència en Biotecnologia de la 476Generalitat de Catalunya. We thank technical support from CCiT-UB. 477 41 21 42 478Figure 1. Degradation pathways of 1,1,2-TCA: (a) dehydrochlorination, (b) 479hydrogenolysis, (c) dichloroelimination, (d) hydrolysis. Numbers indicate the 480dechlorinating agent: (1) base mediated abiotic reaction in aqueous solution (Pagana et 481al. 1998); (2) Desulfitobacterium sp. strain PR (Zhao et al. 2015); (3) Dehalobacter and 482Dehalogenimonas spp (Grostern and Edwards, 2006; Mortan et al. 2017, Yan et al. 4832009); (4) nanosized zero-valent iron (Song and Carraway, 2005); (5) zero valent iron 484and zinc (Patterson et al. 2016); (6) Pseudomonas sp. (Castro and Belser, 1990). Bold 485arrow: biotic reaction; dashed arrow: abiotic reaction. 486 487 488 43 22 44 489Figure 2. Concentration and carbon isotope patterns of 1,1,2-TCA (circles) and VC 490(triangles) during dichloroelimination of 1,1,2-TCA in a Dehalogenimonas-enrichment 491culture prepared with either acetate (empty symbols) or pyruvate (filled symbols) as 492carbon source. The error bars show the one standard deviation (1σ) for duplicate 493measurements. For isotope values the error bars are smaller than the symbols. The 494average δ13C of 1,1,2-TCA in the controls (dashed line) and models fit to isotope data 495from the substrate (eq 2, black solid line) and product (eq 3, grey solid line) are shown. 496 497 498 499 45 23 46 500Figure 3. Double logarithmic plot according to the Rayleigh equation (eq 2) of the 501carbon (A) and chlorine (B) isotope ratios versus the residual concentration of 1,1,2- 502TCA during dichloroelimination by a Dehalogenimonas-containing culture prepared 503with either acetate (empty symbols) or pyruvate (filled symbols) as carbon source. The 504error bars show the one standard deviation (1σ) for duplicate measurements and doted 505lines represent the 95% C.I. of the linear regression determined by SigmaPlot. 506 507 508 47 24 48 509Figure 4. Dual C-Cl isotope plot during dichloroelimination of 1,1,2-TCA in a 510Dehalogenimonas-containing enrichment culture. The error bars show the one standard 511deviation (1σ) for duplicate measurements. For C isotope values the error bars are 512smaller than the symbols. Doted lines represent the 95% C.I. of the linear regression 513determined by SigmaPlot. 514 515 516 49 25 50 517 Table 1. Comparison of ε and AKIE values for C and Cl isotopes assuming either stepwise or concerted reductive dechlorination of chlorinated 518 ethanes and propanes. AKIEC AKIECl Compound Degradation experiment ɛC (‰) Stepwise Concerted ɛCl (‰) Stepwise Concerted Λ Reference 1,1,2-TCA Dehalogenimonas -containing culture -6.9 ± 0.4 1.0138 ± 0.0008 1.0069 ± 0.0004 -2.7 ± 0.3 1.0082 ± 0.0009 1.0041 ± 0.0005 2.5 ± 0.2 This study 1,1,2-TCA Anoxic microcosms -2.0 ± 0.2 1.0040* 1.0020* n.m. n.m. Hunkeler et al. (2002) −14.6 ± 0.7 1.0301 1.0148 Laboratory column 20% (w/w) 1,1,2-TCA to to to n.m n.m Patterson et al. (2016) Fe(0)/organic carbon amendment −0.7 ± 0.1 1.0014* 1.0007* Abiotic laboratory column with 1,1,2-TCA −12 ± 5 1.0246* 1.0121* n.m n.m Patterson et al. (2016) Fe(0) without organic carbon amendment -13.6 ± 0.5 1.028 ± 0.001 1,1,1-TCA abiotic by Cr(II), Fe(0) and Cu and Fe mix to to n.a n.m. n.m. Elsner et al. (2007) -15.8 ± 0.6 1.033 ± 0.001 1,1,1-TCA abiotic by Fe(0) -7.8 ± 0.4 1.0158 ± 0.0008 n.a -5.2 ±0.2 1.0160 ± 0.0006 1.5 ± 0.1 Palau et al. (2014) abiotic degradation mediated by biotic FeS −10.3 to 1,1,1-TCA n.m. n.m. Broholm et al. (2014) formation in bioaugmented microcosms −14.0 Dehalobacter-containing culture (whole -1.8 ± 0.3 1.0036 ± 0.0006 Sherwood Lollar et al. 1,1,1-TCA n.a n.m. n.m. cell and cell-free extracts) -0.8 ± 0.3 1.0016 ± 0.0006 (2010) Dehalococcoides mccartyi strains (195 and -28.4 ± 3.7 1.059 ± 0.008 1.029 -4.6 ± 0.7 1.009 ± 0.001 1.005 6.9 ± 1.2 1,2-DCA Franke et al. (2017) BTF08) -30.9 ± 3.6 1.066 ± 0.008 1.031 -4.2 ± 0.5 1.009 ± 0.001 1.004 7.1 ± 0.2 Dehalococcoides mccartyi strains (195 and -29.0 ± 3.0 1.062 1.030 1,2-DCA n.m n.m Schmidt et al. (2014) BTF08) -30.8 ± 1.3 1.066 1.033 1,2-DCA Dehalococcoides-containing culture −33.0 ± 0.4 1.0707 ± 0.0009 1.0341 ± 0.0004 −5.1 ± 0.1 1.0051 ± 0.0001* 6.8 ± 0.2 Palau et al. (2017) −12.0 ± 1,2-DCA Dehalogenimonas-containing culture −23 ± 2 1.048 ± 0.004 1.024 ± 0.003 1.0121 ± 0.0008* 1.89 ± 0.02 Palau et al. (2017) 0.8 1,2-DCA Anoxic microcosms -32 ± 1 1.069 ± 0.002* 1.033 ± 0.001* n.m n.m Hunkeler et al. (2002) 1,2-DCA abiotic by Zn(0) -29.7 ± 1.5 1.06 – 1.07 1.03 n.m. n.m. Vanstone et al. (2008) -10.5 ± 0.6 1.021 ± 0.002 Dehalobacter-containing culture (whole Sherwood Lollar et al. 1,1-DCA and and n.m. n.m. cell and cell-free extracts) (2010) -7.9 ± 0.9 1.016 ± 0.002 -17.0 ± 0.6 1.035 ± 0.001 1.0173 ± 0.0006 1,1,2,2-TeCA abiotic by Cr(II), Fe(0) and Cu and Fe mix to to to n.m. n.m. Elsner et al. (2007) -19.3 ± 0.7 1.040 ± 0.001 1.0196 ± 0.0008 1,1,2,2-TeCA Abiotic by Cr(II) -12.7 ± 1.2 1.026 ± 0.001 1.013* n.m n.m Hofstetter et al. (2007) PCA Abiotic by Cr(II) -14.7 ± 0.6 1.0303 ± 0.0006 1.0149* n.m n.m Hofstetter et al. (2007) HCA Abiotic by Cr(II) -10.4 ± 0.5 1.0212 ± 0.0005 1.0105* n.m n.m Hofstetter et al. 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