This is the submitted version of the article: Blázquez Palli, Natàlia; Rosell, Mònica; Varias, Joan (Litoclean, S.L.); [et al.]. «Integrative isotopic and molecular approach for the diagnosis and implemen- tation of an efficient in-situ enhanced biological reductive dechlorination of chlorinated ethenes». Water Research, Vol. 167 (December 2019), art. 115106. DOI 10.1016/j.watres.2019.115106 This version is avaible at https://ddd.uab.cat/record/212958 under the terms of the license 1 Integrative isotopic and molecular approach for the diagnosis and 2 implementation of an efficient in-situ enhanced biological 3 reductive dechlorination of chlorinated ethenes 4 Natàlia Blázquez-Pallía,b, Mònica Rosellc, Joan Variasb, Marçal Boschb, Albert Solerc, 5 Teresa Vicenta, Ernest Marco-Urreaa,* 6 a Departament d’Enginyeria Química, Biològica i Ambiental, Universitat Autònoma de 7 Barcelona (UAB), c/ de les Sitges s/n, 08193 Cerdanyola del Vallès, Spain. b 8 Litoclean, S.L., c/ Numància 36, 08029 Barcelona, Spain. c 9 Grup MAiMA, SGR Mineralogia Aplicada, Geoquímica i Geomicrobiologia, 10 Departament de Mineralogia, Petrologia i Geologia Aplicada, Facultat de Ciències de la 11 Terra, Institut de Recerca de l’Aigua (IdRA), Universitat de Barcelona (UB), c/ Martí 12 Franquès s/n, 08028 Barcelona, Spain. 13 * 14 Corresponding author: 15 Departament d’Enginyeria Química, Biològica i Ambiental, Universitat Autònoma de 16 Barcelona (UAB), c/ de les Sitges s/n, 08193 Cerdanyola del Vallès, Spain. E-mail: 17 ernest.marco@uab.cat, Phone: +34 935812694. 18 19 Total number of pages (including cover): 31 20 Words (including references): 7971 21 Figures: 6 22 Tables: 1 1 23 Abstract 24 Based on the previously observed intrinsic bioremediation potential of a site originally 25 contaminated with perchloroethene (PCE), field-derived lactate-amended microcosms 26 were performed to test different lactate isomers and concentrations, and find clearer 27 isotopic and molecular parameters proving the feasibility of an in-situ enhanced reductive 28 dechlorination (ERD) from PCE-to-ethene (ETH). According to these laboratory results, 29 which confirmed the presence of Dehalococcoides sp. and the vcrA gene, an in-situ ERD 30 pilot test consisting of a single injection of lactate in a monitoring well was performed 31 and monitored for 190 days. The parameters used to follow the performance of the ERD 32 comprised the analysis of i) hydrochemistry, including redox potential (Eh), and the 33 concentrations of redox sensitive species, chlorinated ethenes (CEs), lactate, and acetate; 34 ii) stable isotope composition of carbon of CEs, and sulphur and oxygen of sulphate; and 35 iii) 16S rRNA gene sequencing from groundwater samples. Thus, it was proved that the 36 injection of lactate promoted sulphate-reducing conditions, with the subsequent decrease 37 in Eh, which allowed for the full reductive dechlorination of PCE to ETH in the injection 38 well. The biodegradation of CEs was also confirmed by the enrichment in 13C and carbon 39 isotopic mass balances. The metagenomic results evidenced the shift in the composition 40 of the microbial population towards the predominance of fermentative bacteria. Given the 41 success of the in-situ pilot test, a full-scale ERD with lactate was then implemented at the 42 site. After one year of treatment, PCE and trichloroethene were mostly depleted, whereas 43 vinyl chloride (VC) and ETH were the predominant metabolites. Most importantly, the 44 shift of the carbon isotopic mass balances towards more positive values confirmed the 45 complete reductive dechlorination, including the VC-to-ETH reaction step. The 46 combination of techniques used here provides complementary lines of evidence for the 47 diagnosis of the intrinsic biodegradation potential of a polluted site, but also to monitor 2 48 the progress, identify potential difficulties, and evaluate the success of ERD at the field 49 scale. 50 51 Keywords: Chlorinated ethenes; Organohalide-respiring bacteria; Dehalococcoides; In- 52 situ pilot test; Enhanced reductive dechlorination; Carbon isotopic mass balance. 53 3 54 1. Introduction 55 Chlorinated ethenes (CEs) are among the most ubiquitous anthropogenic groundwater 56 contaminants due to their widespread use in industry and recalcitrance under oxic 57 conditions. CEs are considered priority substances (ATSDR, 2017) and have maximum 58 contaminant levels in groundwater set by the 2008/105/EC European Directive (European 59 Commission, 2008). Perchloroethene (PCE) is resistant to aerobic degradation but under 60 anoxic conditions can undergo reductive dechlorination to the less-chlorinated ethenes 61 trichloroethene (TCE), cis- and trans-dichloroethene (cis-, trans-DCE), vinyl chloride 62 (VC), and the harmless end-product ethene (ETH). 63 Organohalide-respiring bacteria (OHRB) provide a potential solution to detoxify sites 64 impacted with CEs due to their capability to use organochlorides as electron acceptors to 65 support their growth, which result in a stepwise reduction of CEs (Bradley, 2000; Leys et 66 al., 2013). Several OHRB can partially dechlorinate PCE to TCE or cis-DCE, e.g. 67 Geobacter sp., but Dehalococcoides sp. (Dhc) and Dehalogenimonas sp. are the only 68 genera, to date, capable of fully dechlorinating CEs to ETH (Adrian and Löffler, 2016; 69 Yang et al., 2017b). Reductive dehalogenases (rdh) are the key enzymes driving 70 organohalide respiration and can be used as biomarkers to investigate the intrinsic 71 bioremediation potential at polluted sites (Blázquez-Pallí et al., 2019; Dugat-Bony et al., 72 2012; Hermon et al., 2019; Scheutz et al., 2008). Compound-specific stable isotope 73 analysis (CSIA) can provide a complementary line of evidence to monitor biodegradation 74 of chlorinated compounds in groundwater (Aelion et al., 2009; Hermon et al., 2018; 75 Hunkeler et al., 1999; Nijenhuis et al., 2007; Palau et al., 2014). This technique measures 13 76 a specific stable isotope ratio (e.g. C/12C) within molecules and is based on light and 77 heavy isotopes degrading at slightly different rates during biochemical transformations. 78 Such shift in the isotopic composition (δ) of the molecule can be used to confirm and 79 quantify in-situ biodegradation and distinguish degradation pathways (Elsner, 2010). This 4 80 is possible because the isotopic enrichment caused by physical processes such as 81 volatilization, sorption, or dilution, is considered to be negligible (Aelion et al., 2009; 82 Hunkeler et al., 1999). Similarly, the isotopic composition of certain non-halogenated 83 potential electron acceptors can be used to trace changes in the redox potential of the 84 groundwater. For instance, a dual isotope approach can be used to measure the extent of 85 sulphate reduction because this reaction results in an enrichment in both 34S and 18O in 86 the residual sulphate (Wu et al., 2011). 87 When OHRB are present but the electron donor becomes a limiting factor, the 88 groundwater contaminated with CEs can show little to no dechlorination past cis-DCE, 89 resulting in the accumulation of toxic intermediates (DCE or VC stall) (Bradley, 2000; 90 Stroo and Ward, 2010). To avoid this, groundwater can be conditioned with organic 91 fermentable substrates (e.g. lactate), which can generate reducing equivalents that 92 promote the sequential dechlorination to ETH. This bioremediation approach is 93 commonly referred to as biostimulation or ERD, which stands for enhanced reductive 94 dechlorination (Adrian and Löffler, 2016; Leeson et al., 2004). To date, many studies 95 have focused on laboratory methodologies to assess and characterize the intrinsic 96 bioremediation potential of CEs-polluted sites by OHRB (Buchner et al., 2015; Courbet 97 et al., 2011; Ebert et al., 2010; Kuder et al., 2013; Lee et al., 2016; Lu et al., 2009; 98 Matteucci et al., 2015; Nijenhuis et al., 2007; Slater et al., 2001; Tarnawski et al., 2016; 99 Yu et al., 2018), but few have reported results after applying ERD and CSIA at the field 100 scale (Herrero et al., 2019; Hirschorn et al., 2007; Song et al., 2002). 101 The research reported in Blázquez-Pallí et al. (2019) presented a multi-method 102 approach to assess the intrinsic biodegradation potential of an industrial site in Barcelona 103 (Spain) polluted by CEs. Obtained results discouraged natural attenuation as a 104 remediation strategy due to cis-DCE stall, and recommended an ERD injecting lactate as 105 electron donor. In line with those conclusions, the present study aimed at continuing the 5 106 remediation work at that same aquifer and apply an integrated isotopic and molecular 107 approach that provides complementary lines of evidence to implement, monitor and 108 assess an efficient in-situ biodegradation of CEs at the site during an ERD. To this end, a 109 more detailed laboratory study focusing, as well, on the concentration and isomeric form 110 of lactate required for the complete dechlorination of PCE to ETH was performed first, 111 followed by the monitoring of the in-situ ERD pilot test and the final full-scale treatment. 112 The used methodology in all cases combined (1) the acquisition of hydrogeochemical 113 data, (2) microcosms experiments, (3) molecular techniques (i.e. identification of selected 114 biomarkers and 16S rRNA high-throughput sequencing of groundwater samples), and (4) 115 stable isotopes of sulphate (as model redox sensitive species), and carbon from the target 116 contaminants (CEs). 117 118 2. Materials and methods 119 2.1. Chemicals 120 PCE (99.9% purity) was from Panreac; laboratory-grade sodium DL-lactate (≥98% 121 purity) from Sigma-Aldrich (hereinafter, Lactate-1), and food-grade sodium L-lactate 122 (97% purity, at 60% w/w) from Purac (Corbion) (hereinafter, Lactate-2). Other chemicals 123 and reagents used for the present study were purchased from Sigma-Aldrich, 124 Thermofisher and Bio-Rad at scientific grade or higher. 125 2.2. Study site 126 The studied site is located in the Barcelona province (Spain) and the aquifer is mainly 127 constituted by three lithological units: i) a lower layer of reddish clay loams, ii) an 128 intermediate layer of brown and silty clays, and iii) an upper layer of ochre silty clays. 129 Given such lithology, the hydraulic conductivity is generally low, only improved by 130 localized gravel and sandstone areas that increase the permeability of the media and could 131 act as preferential flow paths. A hydrogeological cross section of the site is depicted in 6 132 Figure S1, and a more detailed description of the aquifer can be found elsewhere 133 (Blázquez-Pallí et al., 2019). A preliminary site characterization revealed a significant 134 PCE plume, accompanied by minor amounts of TCE, cis-DCE and VC. Originally, it was 135 treated by a combination of pump and treat (P&T) and dual-phase extraction (DPE, 136 vapour and groundwater). Pumped groundwater was later treated through an air stripping 137 system. Groundwater flowed naturally in the NW–SE direction (Figure 1); however, 138 under such ongoing remediation, all the extraction points were almost dried due to the 139 low productivity of the aquifer. These P&T and DPE systems were halted partially during 140 the in-situ pilot test (only remained active in some wells outside the pilot test area), and 141 completely for the full-scale bioremediation. However, certain influence of this pumping 142 on the groundwater flow direction during the in-situ pilot test cannot be discarded. 143 2.3. Microcosm experiments 144 Groundwater with fine sediments was collected from the intended pilot scale injection 145 well PZ-2 (Figure 1) with a peristaltic pump on October 4th, 2016, in transparent sterile 146 glass bottles that were sealed with PTFE caps as described elsewhere (Blázquez-Pallí et 147 al., 2019). These groundwater samples were kept in the dark at 4ºC until the following 148 day, when microcosms were set up. Four different treatments were prepared, at least in 149 duplicate: (1) control containing only groundwater, (2) groundwater with Lactate-1 at ~ 150 3 mM, (3) groundwater with Lactate-2 at ~ 3 mM, and (4) groundwater with Lactate-2 at 151 ~ 15 mM. Each 100 mL microcosm bottle contained 65 mL of sampled groundwater with 152 fine sediments and the lactate concentrations described above. All microcosms were 153 prepared in an anoxic glovebox and incubated in the dark at 25 °C. Microcosms that fully 154 dechlorinated CEs were reamended with TCE and transferred to sterilized anoxic 155 synthetic medium (3–7% v/v) described elsewhere (Martín-González et al., 2015) during 156 the exponential degradation phase of CEs. 7 157 For the analysis of the site-specific carbon isotopic fractionation (εC) during 158 degradation of PCE, six parallel anoxic cultures were prepared with the abovementioned 159 defined medium and groundwater from PZ-2 (1.5% v/v) as inoculum. Each culture was 160 spiked with PCE (235 µM), and sacrificed with NaOH (10 M) at 0, 1, 6, 77, 88, 96 and 161 97% of PCE degradation. Three different controls were prepared in duplicate: (1) killed 162 controls with PCE, (2) killed controls without PCE, and (3) abiotic controls with PCE. 163 2.4. Implementation of the ERD in-situ pilot test 164 The pilot test consisted of a unique injection of lactate at well PZ-2 (Figure 1) on 165 October 25th, 2016 (t0). The product injected was an aqueous solution of Lactate-2 diluted 166 with groundwater outflowing from the air stripping system installed at the site. The total 167 volume of substrate injected was lower than 10% of the treatment zone volume. The 168 design parameters of the in-situ pilot test were based on the recommendations from 169 Leeson et al. (2004) and Dugat-Bony et al. (2012). 170 2.5. Monitoring of the ERD in-situ pilot test 171 Field parameters (i.e. Eh, pH, T, and electric conductivity (EC)) were measured in-situ 172 and groundwater was collected with a peristaltic pump from the injection well PZ-2 and 173 nearby wells PZ-1, PZ-3, PZ-6, PZ-22, MW-6, and Prof.D (Figure 1) as described 174 elsewhere (Blázquez-Pallí et al., 2019). Sampling campaigns were carried out the day 175 before injection of lactate (t-1, October 24th, 2016) and the next 2, 9, 20, 50, 86, 142 and 176 190 days after the injection (hereinafter, ti). Samples collected for both CSIA and CEs 177 concentration were immediately killed with NaOH (pH>10) and stored at 4ºC until 178 analysed. Short-chain volatile fatty acids (VFAs) were analysed from groundwater 179 samples filtered on site (0.20 µm) and stored in borosilicate tubes at 4ºC until analysed. 180 2.6. Isotopic evaluation of the full-scale ERD treatment 181 A full-scale ERD treatment was implemented by Litoclean, S.L. on August 2017. The 182 bioremediation strategy consisted in the injection of Lactate-2 every three months for the 8 183 period of a year. Between 30 and 50 out of a total of 66 monitoring wells at the site were 184 used for the injection of the substrate at every event but changing the distribution of the 185 injections to ensure the maximum coverage of the plume area. On September 2018, after 186 one year of treatment, groundwater was sampled from wells PZ-3, PZ-5, PZ-22, MW-3, 187 MW-6, and MW-7 (Figure 1) following the same methodology mentioned in section 2.5. 188 Samples were analysed for CEs concentrations and CSIA, and data were compared to 189 previous (before the in-situ pilot test) chemical and isotopic values for each well. 190 2.7. Analytical methods 191 CEs concentrations were analysed from 500 µL headspace samples by gas 192 chromatography (GC) coupled to a flame ionization detector (FID) as reported by Martín- 193 González et al. (2015). VFAs (lactate, pyruvate, acetate, formate) were analysed by high 194 performance liquid chromatography (HPLC) from filtered liquid samples as described 195 elsewhere (Mortan et al., 2017). 196 Stable carbon isotopes of CEs were analysed with an Agilent 6890 GC coupled to an 197 IRMS at Centres Científics i Tecnològics de la Universitat de Barcelona (CCiT-UB), 198 following the procedure described in Blázquez-Pallí et al. (2019). Carbon isotopic 199 compositions in all samples are presented in delta notation (δ13C, in ‰), relative to the 200 international standard Vienna Pee Dee Belemnite (VPDB), following 𝑅𝑠𝑎𝑚𝑝𝑙𝑒 201 𝛿 13 𝐶 = ( − 1) · 1000 (1) 𝑅𝑠𝑡𝑑 13 202 where Rsample and Rstd represent the isotope ratios (e.g. C/12C) of the sample and the 203 standard, respectively (Elsner, 2010). Instrument uncertainty was considered as the 204 standard deviation (1σ) of duplicate measurements. For field data, the degradation is 205 considered significant if the shift in δ13C is >2‰ compared to its original value (Hunkeler 206 et al., 2008). 9 207 Carbon isotopic mass balance for CEs at each monitoring well was calculated 208 according to Aeppli et al. (2010) and Hunkeler et al. (1999), as follows 209 𝛿 13 𝐶𝑠𝑢𝑚 = 𝑥𝑃𝐶𝐸 · 𝛿 13 𝐶𝑃𝐶𝐸 + 𝑥𝑇𝐶𝐸 · 𝛿 13 𝐶𝑇𝐶𝐸 + 𝑥𝐷𝐶𝐸 · 𝛿 13 𝐶𝐷𝐶𝐸 + 𝑥𝑉𝐶 · 𝛿 13 𝐶𝑉𝐶 (2) 210 where x is the molar fraction of each substance with respect to the total molar mass (sum 211 of CEs for which δ13C values were available) for each sample at each sampling event. In 212 this case, as PCE was the only contaminant spilled in this aquifer, the δ13Csum must remain 213 constant as long as (1) PCE released along the time and space had the same isotopic 214 composition, (2) the unique transformation pathway was reductive dechlorination, (3) VC 215 does not further degrade to ETH when considered in the balance (Aeppli et al., 2010; 216 Hunkeler et al., 1999; Palau et al., 2014). 217 The logarithmic form of the simplified Rayleigh equation correlates changes in the 218 carbon isotope ratios (Rt/R0) and changes in concentrations (f=Ct/C0) with time for a 219 closed system (Elsner, 2010), and the obtained epsilon (εC) represents the carbon isotopic 220 fractionation, as follows 𝑅𝑡 221 ln ( ) = 𝜀𝐶 · ln(𝑓) (3) 𝑅0 222 where Rt/R0 can be expressed as (δ13Ct + 1) / (δ13C0 + 1). 223 The analysis of major anions (HCO3-, NO3-, Cl-, SO4-2) and cations (Na+, K+, Ca+2, 224 Mg+2) was performed at CCiT-UB. Total concentrations of Na+, K+, Ca+2, Mg+2 were 225 analysed by inductively coupled plasma-optic emission spectrometry (ICP-OES, Optima 226 3200 RL) and by inductively coupled plasma mass spectrometry (ICP-MS, Elan 6000). 227 NO3-, Cl-, and SO4-2 concentrations were determined by HPLC using a WATERS 515 228 HPLC pump with an IC-PAC anion column and a WATERS detector (mod 432), while 229 HCO3- was measured by titration (METROHM 702SM Titrino). The predominant 10 230 equilibrium systems controlling the Eh were investigated via Eh–pH predominance 231 diagrams prepared with the MEDUSA code (Puigdomènech, 2010). 232 Dissolved SO4-2 detected in groundwater was precipitated as BaSO4 as reported 233 elsewhere (Dogramaci et al., 2001) and its sulphur and oxygen isotopic compositions 234 were analysed following Rodríguez-Fernández et al. (2018). Results are presented in delta 235 notation (δ34S-SO4-2 and δ18O-SO4-2, in ‰), relative to the international standards, Vienna 236 Standard Mean Oceanic Water (VSMOW) for δ18O and Vienna Canyon Diablo Troillite 237 (VCDT) for δ34S. 238 2.8. DNA extraction, PCR and 16S rRNA gene high-throughput sequencing 239 DNA for molecular analyses was extracted from TCE-enriched field-derived cultures 240 (three transfers into fresh synthetic medium) that were originally amended with Lactate- 241 2 (~ 3 mM) and from well PZ-2. Cell harvesting, genomic DNA isolation and PCR 242 reaction for amplification of the vcrA gene were performed as described in Blázquez-Pallí 243 et al. (2019). Primer sets used are detailed in Table S1. 244 To investigate the OHRB involved in the biodegradation of CEs to ETH, the dilution- 245 to-extinction method (Löffler et al., 2005) was applied in 20-mL vials containing 12 mL 246 of the anoxic synthetic medium mentioned in section 2.3, but using 1 mL of active TCE- 247 enriched field-derived culture as inoculum and cis-DCE as electron acceptor. After three 248 extinction series, the more diluted vial showing ETH formation was used as inoculum for 249 serum bottle microcosms and, after consuming 10 µM cis-DCE, was selected for 16S 250 rRNA analysis. After DNA extraction, amplicons of the region V3–V4 for 16S rRNA 251 genes were amplified with primers S-D-Bact-0341-b-S-17 and S-D-Bact-0785-a-A-21 252 (Klindworth et al., 2013) with the Illumina MiSeq sequencing platform at Serveis de 253 Genòmica i Bioinformàtica from Universitat Autònoma de Barcelona (Spain). 254 To characterize the impact of the injection of lactate on the bacterial community 11 255 structure after the in-situ pilot test, 80 mL of groundwater were collected from wells PZ- 256 2, PZ-1, and PZ-22 at t-1 and t142. Samples were stored at -20ºC until DNA extraction and 257 16S rRNA gene high-throughput sequencing were performed at AllGenetics & Biology 258 (A Coruña, Spain). For DNA extraction, samples were centrifuged at 5000 g for 1 h and 259 the pellet was transferred to PowerBead tubes of the DNeasy Powersoil DNA isolation 260 kit (Qiagen). DNA was isolated following the instructions of the manufacturer. For library 261 preparation, a fragment of the bacterial 16S rRNA region of around 450 bp was amplified 262 using primers Bakt_341F and Bakt_805R (Herlemann et al., 2011). The pool was 263 sequenced in a MiSeq PE300 run (Illumina). 264 265 3. Results and discussion 266 3.1. Feasibility and markers for complete ERD at laboratory scale 267 3.1.1. Assessment of the lactate isomers and concentration in microcosm experiments 268 Microcosms were prepared to test whether reductive dechlorination of CEs to ETH 269 was feasible in the injection well PZ-2. Besides the control, which accounted for 270 monitored natural attenuation (MNA), two different isomeric forms of sodium lactate 271 (DL-/Lactate-1 and L-/Lactate-2) were used to test their effect on the lactate fermentation 272 potential of the native microbial community. Moreover, Lactate-2 was used because it 273 was a candidate product in a foreseeable in-situ pilot test. Accordingly, Lactate-2 was 274 amended at two different concentrations to discard inhibition effects. On the one hand, 275 the unamended controls transformed PCE to cis-DCE and VC within 15 days, but ETH 276 was not detected. When microcosms were reamended with TCE (35 µM) at day 25, 277 dechlorination barely passed VC, and ETH was detected at low concentration after 50 278 days (Figure S2A). On the other hand, lactate-amended treatments fully dechlorinated 279 PCE to ETH within 15 days and exhibited similar rates independently of the isomeric 280 form and concentration (Figure S2, B–D). After adding TCE in all three of them at days 12 281 25 and 45, the dechlorination remained active with the consequent accumulation of ETH 282 (Figure S2, B–D). In all amended microcosms, lactate was totally consumed and acetate 283 was produced within 10 days (data not shown). However, excess of electron donor and, 284 thus, of H2, did promote methanogenesis, resulting in a vigorous formation of methane in 285 the microcosm with ~ 15 mM of lactate (Figure S3), which has been reported before 286 (Blázquez-Pallí et al., 2019; Leeson et al., 2004). 287 3.1.2. Identification of vcrA gene and CEs dechlorinating bacteria 288 PCR amplifications with the vcrA gene-targeted primer were run to investigate 289 whether the vcrA gene was present in the TCE-enriched field-derived cultures from well 290 PZ-2 (section 2.8). After gel electrophoresis, observed diagnostic amplicons indicated 291 that the culture contained vcrA gene, which is implicated in the VC-to-ETH 292 dechlorination step (Figure S4). 293 In these TCE-enriched cultures, three dilution-to-extinction series were applied using 294 cis-DCE as electron acceptor to get insight into the OHRB responsible for ETH 295 generation. The 16S rRNA gene amplicon sequencing and the taxonomic assignments 296 revealed that the four predominant genera were Pleomorphomonas (50%), described as 297 nitrogen-fixing bacteria (Im et al., 2006; Madhaiyan et al., 2013; Xie and Yokota, 2005); 298 Desulfomicrobium (18%), which are sulphate-reducing bacteria (Sharak Genthner et al., 299 1997, 1994); and the OHRB Geobacter (14%) and Dehalococcoides (3%) (Figure 2). 300 Geobacter sp. can derive energy from acetate oxidation coupled to PCE-to-cis-DCE 301 dechlorination, but it is also capable of growing on a wide range of non-halogenated 302 electron acceptors (Atashgahi et al., 2016; Sanford et al., 2016). The presence of 303 Dehalococcoides was consistent with the detection of vcrA reductive dehalogenase gene 304 (Figure S4) and ETH generation in lactate-amended microcosms (Figure S2). 305 13 306 3.1.3. Site-specific εC for PCE degradation 307 The analysis of site-specific εC can help estimate the extent of the in-situ 308 biodegradation of contaminants in the field (Elsner, 2010). PCE was depleted in the 309 microcosms inoculated with groundwater from PZ-2 (Figure S5A) but dechlorination was 310 not accompanied by a significant change in its isotopic composition (δ13CPCE) (Figure 311 S5B). In more detail, δ13CPCE shifted only 0.81‰ and 1σ for duplicate measurements 312 were, for all samples, below total instrumental uncertainty of 0.5‰ (Sherwood Lollar et 313 al., 2007). The dechlorination reaction did not fit the Rayleigh model (Eq. 3, R2 = 0.63, 314 εC = -0.2 ± 0.2‰), which deemed this site-specific εC as not significant (ns). In contrast, 315 a stronger enrichment in 13C was observed when the produced TCE was further degraded 316 to cis-DCE (Figure S5B). This non-linear εC for PCE is, most likely, the combined effect 317 of several OHRB simultaneously transforming PCE, with a major contribution of the non- 318 fractionating species. This agrees with the predominance of Geobacter sp. in the cultures 319 (Figure 2), which is reported to have a non-significant εC for PCE and could have lead 320 the PCE-to-TCE dechlorination (Table S2) (Cichocka et al., 2008). Abiotic and NaOH- 321 killed controls did not show PCE losses or degradation as its concentration did not vary 322 significantly throughout the whole experiment (236 ± 8 μM, n=6, Figure S5A) and no 323 degradation products were detected. 324 3.2. In-situ ERD pilot test with lactate at the injection well PZ-2 325 Microcosms of PZ-2, which mimicked the natural attenuation conditions at the site, 326 indicated that reductive dechlorination could stall at an intermediate stage, resulting in an 327 accumulation of cis-DCE and VC (Figure S2A). However, the amendment of lactate 328 resulted in a faster dechlorination of PCE and concomitant generation of ETH (Figure S2, 329 B–D). Considering these results, an in-situ ERD pilot test with Lactate-2 was 330 implemented at well PZ-2. 14 331 3.2.1. Hydrochemistry changes induced by the lactate injection 332 The concentration of major anions (SO4-2, NO3-, HCO3-, Cl-) and cations (Na+, K+, 333 Ca+2, Mg+2) was analysed at t-1, t20, t86 and t190. At the injection well PZ-2, NO3- and SO4- 2 334 were depleted after the addition of the Lactate-2 solution (Figure S6A), whereas Na+, 335 Ca+2, Mg+2 and electric conductivity (EC) increased, responding to the injected electron 336 donor (Table S3). At the monitoring wells PZ-1, PZ-6 and MW-6, NO3- concentrations 337 were depleted as well, but different trends were observed for SO4-2 (Figure S6, A and B). 338 In contrast, no significant changes were observed for those anions in monitoring wells 339 PZ-3, PZ-22, and Prof.D (Figure S6C). 340 Geochemical modelling shows that the aquifer was at nitrate-reducing conditions 341 before injection and that the addition of lactate promoted a shift towards sulphate- 342 reducing and methanogenic conditions (Figure S7). This shift was more extreme in the 343 injection well PZ-2 (at t20) as well as in monitoring wells PZ-1 (at t86), and PZ-6 and MW- 344 6 (both at t190), which agrees with the decrease observed in NO3- and SO4-2 concentrations 345 (Figure S6). In addition, results suggested that the system was controlled by calcite 346 (CaCO3(s)) (Figure S7B), as the pH was maintained within the range of 6–8, the optimal 347 for OHRB (Yang et al., 2017a). 348 δ34S and δ18O values of dissolved SO4-2 from t-1 at each monitoring well were 349 compared with the ones obtained after the lactate injection (t20, t86 or t190). Before the 350 injection, the S and O isotopic composition of SO4-2 at the site showed diverse values 351 corresponding to a mixture among geogenic composition (Triassic or Tertiary recycled 352 gypsum) in well Prof.D, synthetic fertilizers, and an unknown source (with the lowest 34S 353 and 18O values) that could be related to agricultural uses of manure (Otero et al., 2008). 354 This variability points out to a heterogeneity in the origin of the dissolved sulphate in the 355 aquifer (Figure 3). In any case, in the injection well PZ-2, isotopic compositions at t20 356 became significantly more enriched with a shift of Δδ34S = +3.0‰ and Δδ18O = +3.2‰ 15 357 resulting in a slope of 1.1 (Figure 3). These δ34S and δ18O values of sulphate seem 358 influenced by the mixing with the sulphate isotopic composition of the injected water 359 (Figure 3). Unluckily, no values could be obtained for t86 or t190 due to concentrations 360 being below the limit of quantification. An even larger shift was measured in monitoring 361 wells PZ-6 and MW-6 because the slower decrease of SO4-2 concentrations allowed the 362 measurement of more points. By t190, PZ-6 showed the most 34S enriched value measured 363 at the site, also with respect to its t-1 (Δδ34S = +14.4‰), but with a small δ18O enrichment 364 (Δδ18O = +1.0‰) (Figure 3). For MW-6, the most enriched δ34S was measured in t86 365 (Δδ34S = +7.4‰), while the largest Δδ18O, of +4.7‰, was measured in t190 (Figure 3). In 366 contrast, Δδ34S and Δδ18O in monitoring wells PZ-1, PZ-3, PZ-22, and Prof.D were not 367 significant (Figure 3). A wide variation in the dual S–O slopes have been described for 368 bacterial sulphate reduction in different natural environments (from 0.23 to 1.7) (Antler 369 et al., 2013; Mizutani and Rafter, 1973). These differences have been linked to the net 370 sulphate reduction rate and to the recycling of intermediate species (such as sulphite that 371 facilitate oxygen isotope exchange with H2O) back to sulphate. In environments where 372 sulphate reduction is fast, this sulphite re-oxidation is minimal resulting in lower slopes 373 due to sulphur isotopes increasing faster than oxygen isotopes (Antler et al., 2013). Thus, 374 obtained results (with the available data points) proved sulphate reduction after lactate 375 injection only in PZ-6 and MW-6 by isotopic enrichment rather than pure mixing with 376 injected water. 377 3.2.2. Enhanced biodegradation of chlorinated ethenes at the injection well PZ-2 378 PCE was the main toxic substance dissolved in groundwater of PZ-2 before lactate 379 injection (Figure 4A). After biostimulation, Eh decreased from +300-400 mV to ~ -50 380 mV by t20 (Figure 4D) and acetate was detected for the first time by t50 (Figure 4C). The 381 generation of reducing equivalents from lactate fermentation and the subsequent dramatic 382 decrease of Eh favoured dechlorination past cis-DCE and VC up to ETH by t190 (Figure 16 383 4A). The full dechlorination of CEs to ETH by t190 was also confirmed by the enrichment 384 in the isotopic mass balance (δ13Csum), which changed from -33 ± 2‰ to -9.0 ± 0.5‰ 385 (Figure 4B). This difference in the δ13Csum is due to not considering the isotopic 386 composition of ETH after VC degradation (which would be depleted as lighter isotopes 387 react faster). Hence, δ13Csum became less negative as VC to ETH reaction progressed. In 388 detail, the PCE to TCE reaction at PZ-2 did not change δ13CPCE significantly and remained 389 constant throughout the whole monitoring period (Figure 4B), emulating the results of 390 the microcosm experiments (εC = ns) (section 3.1.3.). Conversely, TCE and cis-DCE 391 showed variations in δ13C during the in-situ dechlorination process, which were up to 392 +7‰ for TCE and +16‰ for cis-DCE at the measured times (Figure 4B). Lastly, VC 393 could be nicely traced, exhibiting a remarkable variation of δ13C (shift up to +49‰) 394 including its formation from cis-DCE and degradation to ETH by t190 (Figure 4B). 395 3.2.3. Impact of the in-situ ERD pilot test at wells within the direct radius of influence 396 Lactate and/or acetate were progressively detected in PZ-1 (at t2), PZ-6 (at t20) and 397 finally in MW-6 (at t142), indicating the arrival of the injected solution through the 398 preferential groundwater flow paths (with a stronger N–S direction than at natural 399 conditions, Figure 1), thus drawing the direct radius of influence of the in-situ ERD pilot 400 test. Another evidence of the affectation at these wells was the dramatic decrease in Eh 401 like the one observed in the injection well PZ-2 (Figures S8, S9, S10). However, even 402 when reaching similar negative Eh values (down to -100 mV), the PCE reductive 403 dechlorination mainly stalled at cis-DCE (most evident in PZ-1) (Figures 5, S8, S9, S10), 404 while in PZ-6 (farther from PZ-2) cis-DCE started to pass to VC and ETH by t190 (Figure 405 S9A). These concentration data agree with δ13Csum of CEs remaining constant in the three 406 wells (~ -32‰ over 190 days, Figures 5, S8B, S9B, S10B), so there was no significant 407 alternative degradation pathways besides ERD, and ETH generation was still minor 408 compared to PZ-2. Curiously, the most inefficient ERD (cis-DCE stall confirmed by no 17 409 changes in its δ13Ccis-DCE after t50) of these wells occurred in PZ-1, where the initial drop 410 in the Eh was followed by a rebound to more positive values after t86 (Figure S8D). In 411 fact, as said before, according to sulphate isotopic composition, PZ-1 did not reach 412 significant sulphate reducing conditions (Figure 3). This was most likely caused by the 413 consumption and absence of electron donor from t86 onwards or fast entrance of oxidizing 414 groundwater. In contrast, possibly due to the combination of preferential flow paths and 415 the influence of external active remediation systems (P&T and DPE) on the groundwater 416 flow direction, certain amounts of organic acids arrived to wells PZ-6 and MW-6 (Figure 417 S9C, S10C), which were the farthest from PZ-2. Their Eh decreased much more 418 progressively after t50, and remained negative until t190, allowing sulphate reducing 419 conditions and further degradation to VC (Figure S9D, S10D). 420 421 3.2.4. Impact of the in-situ pilot test in wells outside the direct radius of influence 422 Lactate and acetate were not detected in wells PZ-3, PZ-22, and Prof.D (Figures S11C, 423 S12C and S13C). Accordingly, Eh did not reach negative values in any of them (Figures 424 S11D, S12D and S13D), not allowing to prove a direct effect of the ERD pilot test. In 425 well PZ-3, Eh decreased to +60 mV at t86 but it rebounded to initial conditions at t190 426 (Figure S11D). Curiously, CEs concentrations decreased globally in PZ-3 and PZ-22, but 427 PCE molar fraction increased in respect to its daughter products, which could indicate a 428 solubilization and dilution of the PCE adsorbed to the soil due to the injected volume 429 (Figure S14). Therefore, the 13C enrichments observed for TCE and cis-DCE in PZ-3 (by 430 +11.5‰ and +8.1‰, respectively, Figure S11) could be attributed to reductive 431 dechlorination in the well or, most probably, associated to the dispersion from upstream 432 contaminants (Figure 1). In wells PZ-22 and Prof.D, changes in CEs concentrations 433 (Figures S12A and S13A) were not associated to significant shifts in 13C values (Figures 434 S12B and S13B), confirming this dilution effect. 18 435 3.2.5. Changes in the native microbial community induced by the in-situ pilot test 436 The effect of lactate injection on the microbial community of the aquifer was 437 investigated by high-throughput sequencing of the 16S rRNA region of selected 438 groundwater samples. PZ-2, PZ-1 and PZ-22 at t-1 and t142 were chosen based on the 439 results obtained in the pilot test, as they represented three different scenarios: i) PCE-to- 440 ETH reaction at the injection well (PZ-2), ii) cis-DCE stall (PZ-1), and iii) a well not 441 directly impacted by the injection of lactate (PZ-22). In PZ-2, the most abundant phyla at 442 t-1 were Proteobacteria (58%), Planctomycetes (11%), Bacteroidetes (6%), 443 Verrucomicrobia (6%), Chlamydiae (4%), Actinobacteria (4%), and Acidobacteria (2%), 444 whereas at t142, the community shifted to Firmicutes (67%), Bacteroidetes (14%), 445 Proteobacteria (14%), and Tenericutes (2%) (Figure 6, Table S4). Within Firmicutes, 446 bacteria from the family Veillonellaceae (26%), and Erysipelotrichaceae (9%) were the 447 most abundant; and the same occurred for the genus Desulfovibrio (3%), from the 448 Proteobacteria phylum (data not shown). The microbial composition after the injection 449 of lactate was dominated by fermenting bacteria that were probably induced by the 450 injection of lactate (Fennell et al., 1997; He et al., 2007; Tegtmeier et al., 2016). In 451 contrast, the abundance of the Firmicutes phylum was very limited in well PZ-1 452 (abundance of 4% at t142) and almost insignificant in well PZ-22, which agrees with the 453 chemical and isotopic results discussed before. 454 3.3. Isotopic evaluation of the full-scale ERD with lactate 455 After the in-situ ERD pilot test, which finished in June 2017, a full-scale ERD with 456 Lactate-2 was implemented on August 2017. After one year of full-scale treatment (from 457 August 2017 to September 2018), an isotopic mass balance of CEs was calculated from 458 6 selected monitoring wells (PZ-3, PZ-5, PZ-22, MW-3, MW-6, MW-7, Figure 1) to 459 evaluate the extent and success of the full-scale ERD. Results revealed that PCE and TCE 460 were completely depleted in 5 out of the 6 monitoring wells and VC and ETH were the 19 461 major end products (Table 1). Originally, δ13Csum at the site was, in average, of -30 ± 2‰, 462 which responded to the original δ13C value of PCE (Blázquez-Pallí et al., 2019). After 463 one year of full-scale ERD, the δ13Csum values in those 5 wells were enriched (ranging 464 from -23.2 ± 0.5‰ to -13 ± 1‰) with respect to the initial value (Table 1). This trend in 465 δ13Csum of becoming more positive was also observed in well PZ-2 during the in-situ pilot 466 test and responds to ETH production. In the case of MW-7, the δ13Csum was already 467 different from the rest of the wells in the characterization campaign (May 2016) 468 (Blázquez-Pallí et al., 2019), but biodegradation up to VC has been also observed after 469 the full-scale (Table 1). Therefore, these results confirmed that the reductive 470 dechlorination of CEs was occurring across the site, albeit at different extent in each well. 471 472 4. Conclusions 473 The present study aimed at validating a multidisciplinary methodology to assess and 474 monitor biodegradation of CEs at the field scale. In this work, hydrochemical, isotopic 475 and microbiological data from the field was examined together with aquifer-derived 476 laboratory microcosms to get an insight into the idiosyncrasies of the aquifer. The first 477 part of this study focused on the diagnosis of the in-situ biodegradation potential of the 478 site for full reductive dechlorination of CEs. The results obtained provided evidence that 479 the stall at cis-DCE and VC was due to the lack of electron donors. Afterwards, the 480 injection of lactate in a monitoring well evidenced that full reductive dechlorination to 481 ETH was achieved within 190 days, and that the reaction followed a different pace in the 482 surrounding monitoring wells. Therefore, the hydrodynamic characteristics of the aquifer 483 possibly dominated the distribution of electron donor, hence, affecting the outcome of the 484 biodegradation reaction in the studied surrounding wells. The methodology proposed here 485 can be easily integrated as a tool in the early stages of site investigation to decide which 486 bioremediation strategy is more suitable according to the specific characteristics of the 20 487 aquifer. Moreover, it can reduce the uncertainties regarding the ongoing processes that 488 are occurring in groundwater systems. Finally, the initial isotopic characterization of the 489 site allowed for applying isotope-mass balance calculations to unequivocally demonstrate 490 ETH generation after the full-scale enhanced reductive dechlorination via lactate 491 injection. This stable isotope-based technique can overcome the difficulties related to the 492 traditional mass balance approaches which are often hindered by natural processes (e.g. 493 dilution, sorption, volatilization). 494 495 5. Acknowledgements 496 This research has been supported by Agencia Estatal de Investigación (CTM2016- 497 75587-C2-1-R, CGL2014-57215-C4-1-R and CGL2017-87216-C4-1-R projects) and co- 498 financed by the European Union through the European Regional Development Fund 499 (ERDF). This work was also partly supported by Generalitat de Catalunya through the 500 consolidate research groups (2017-SGR-14 and 2017SGR-1733) and the Industrial 501 Doctorate grant (2015-DI-064) of N. Blázquez-Pallí. M. Rosell acknowledges a Ramón y 502 Cajal contract (RYC-2012-11920) from MINECO. The Departament d’Enginyeria 503 Química, Biològica i Ambiental of Universitat Autònoma de Barcelona is a member of 504 Xarxa de Referència en Biotecnologia de la Generalitat de Catalunya. We thank Dr. 505 Ivonne Nijenhuis for providing Dehalococcoides mccartyi BTF08, Dr. Cristina 506 Domènech for the geochemical modelling advice, David Fernández-Verdejo for his 507 contribution in the laboratory work, and CCiT-UB and Dr. Roger Puig for excellent 508 technical assistance. 509 510 511 21 512 Figure 1. Water table (blue lines, in m.a.s.l.) representative of the natural groundwater 513 flow at the site and location of monitoring wells. Red numerical codes refer to the wells 514 monitored during the in-situ ERD pilot test. 515 516 517 518 519 520 521 522 22 523 Figure 2. Microbial population in enriched cultures derived from microcosms amended 524 with Lactate-2 from well PZ-2. Average composition obtained from two cultures (A and 525 B) that were enriched with TCE first, and later with cis-DCE for several consecutive 526 dilution series. Total sequence reads for the genera Geobacter and Dehalococcoides were, 527 respectively, 24232/140088 and 5393/140088 in replicate A, and 10421/103805 and 528 3023/103805 in replicate B. Genera with abundance <1% are grouped in “Other”. 529 530 531 532 533 23 534 Figure 3. Dual S–O isotopic compositions of SO4-2 during the in-situ ERD pilot test in 535 all monitored wells. “Injected water” refers to the groundwater collected from across the 536 aquifer, treated through an air stripping system, later mixed with Lactate-2, and injected 537 into well PZ-2 for the pilot test. Samples t86 and t190 for PZ-2 had too low SO4-2 538 concentration for isotopic measurements, while sample t86 for PZ-6 was lost. Dashed 539 black lines represent the slopes estimated for each well with the available values. Shapes 540 of symbols indicate different sampling events, while colours indicate different monitoring 541 wells. 542 543 544 24 545 Figure 4. Time-course of monitored parameters at well PZ-2 during the in-situ pilot test: 546 (A) concentration of CEs; (B) carbon isotopic composition and balance; (C) concentration 547 of organic acids, and (D) redox potential (Eh). Data presented is an average of each 548 sampling event and includes error bars showing the one standard deviation (1σ) for 549 duplicate measurements. 550 551 552 553 554 25 555 Figure 5. Location of the lactate injection well (PZ-2) and the monitoring wells of the in- 556 situ ERD pilot test where the influence of the biostimulation was proved along 190 days, 557 indicating the molar fraction shift of CEs + ETH and the final carbon isotopic mass 558 balance (δ13Csum,t190). The size of the pie charts is related to the total concentration of CEs 559 + ETH in each well and sampling time. 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