journal of environmental sciences 118 (2022) 204–213 Available online at www.sciencedirect.com www.elsevier.com/locate/jes Assessment of aerobic biodegradation of lower-chlorinated benzenes in contaminated groundwater using field-derived microcosms and compound-specific carbon isotope fractionation Alba Trueba-Santiso 1, Jordi Palau 2, Jesica M. Soder-Walz 1, Teresa Vicent 1, Ernest Marco-Urrea 1,∗ 1 Department of Chemical, Biological and Environmental Engineering, Universitat Autònoma de Barcelona (UAB), 08193 Bellaterra, Spain 2 MAiMA group, SGR Applied Mineralogy, Geochemistry and Geomicrobiology, Department of Mineralogy, Petrology and Applied Geology, Faculty of Earth Sciences, Universitat de Barcelona (UB), Martí Franquès s/n, 08028 Barcelona, Spain. a r t i c l e i n f o a b s t r a c t Article history: Biodegradation of lower chlorinated benzenes (tri-, di- and monochlorobenzene) was Received 7 October 2021 assessed at a coastal aquifer contaminated with multiple chlorinated aromatic hydro- Revised 20 December 2021 carbons. Field-derived microcosms, established with groundwater from the source zone Accepted 20 December 2021 and amended with a mixture of lower chlorinated benzenes, evidenced biodegradation Available online 3 January 2022 of monochlorobenzene (MCB) and 1,4-dichlorobenzene (1,4-DCB) in aerobic microcosms, whereas the addition of lactate in anaerobic microcosms did not enhance anaerobic reduc- Keywords: tive dechlorination. Aerobic microcosms established with groundwater from the plume con- Groundwater sumed several doses of MCB and concomitantly degraded the three isomers of dichloroben- Isotope fractionation zene with no observable inhibitory effect. In the light of these results, we assessed the ap- Chlorobenzenes plicability of compound stable isotope analysis to monitor a potential aerobic remediation Aerobic biodegradation treatment of MCB and 1,4-DCB in this site. The carbon isotopic fractionation factors (ε) ob- Anaerobic reductive dehalogenation tained from field-derived microcosms were -0.7‰ ± 0.1 ‰ and -1.0‰ ± 0.2 ‰ for MCB and Natural attenuation 1,4-DCB, respectively. For 1,4-DCB, the carbon isotope fractionation during aerobic biodegra- dation was reported for the first time. The weak carbon isotope fractionation values for the aerobic pathway would only allow tracing of in situ degradation in aquifer parts with high extent of biodegradation. However, based on the carbon isotope effects measured in this and previous studies, relatively high carbon isotope shifts (i.e., ࢞δ 13 C > 4.0 ‰) of MCB or 1,4- DCB in contaminated groundwater would suggest that their biodegradation is controlled by anaerobic reductive dechlorination. © 2022 The Research Center for Eco-Environmental Sciences, Chinese Academy of Sciences. Published by Elsevier B.V. This is an open access article under the CC BY-NC-ND license (http://creativecommons.org/licenses/by-nc-nd/4.0/) ∗ Corresponding author. E-mail: ernest.marco@uab.es (E. Marco-Urrea). https://doi.org/10.1016/j.jes.2021.12.025 1001-0742/© 2022 The Research Center for Eco-Environmental Sciences, Chinese Academy of Sciences. Published by Elsevier B.V. This is an open access article under the CC BY-NC-ND license (http://creativecommons.org/licenses/by-nc-nd/4.0/) journal of environmental sciences 118 (2022) 204–213 205 Thus, engineered bioremediation systems can exploit the Introduction degradation capability of these microorganisms to remediate sites impacted with chlorinated benzenes. A crucial issue in Chlorinated benzenes are extensively used worldwide as in- the implementation of such techniques is the selection of a termediates for the synthesis of pesticides and other chemi- proper tool to monitor the extent of biodegradation. One of cals such as dyes, drugs and plastic materials (Field and Sierra- the most promising tools for both qualitative and quantita- Alvarez, 2008). These compounds are frequently detected in tive estimation of biodegradation processes at contaminated the subsurface nearby industrial areas, as result of its im- field sites is compound-specific stable isotope analysis (CSIA), proper disposal or accidental spills, and they are included in that has been applied successfully in many studies over the the list of priority pollutants in the 2019 US EPA’s Priority List last years (Blázquez-Pallí et al., 2019b; Palau et al., 2017). Typi- of Hazardous Substances (ATSDR, 2019). Their widespread de- cally, molecules containing light isotopes (e.g. 12 C) react faster tection in contaminated sites has led to increased public con- compared to those with heavy isotopes (e.g. 13 C), resulting cern and the need for developing efficient remediation strate- in an enrichment of the heavy isotope in the residual frac- gies. tion of the contaminant, so called kinetic isotopic fractiona- When leaching into groundwater, chlorinated benzenes tion. Nondegradative processes such as dilution, sorption and can undergo both aerobic and anaerobic biodegradation pro- dispersion in the aquifer are not, or only to a much lower cesses, depending on the number and position of chlo- extent, subject to isotope fractionation (Nijenhuis and Rich- rine substituents on the benzene ring and the redox con- now, 2016). Thus, for a given contaminant and reaction mech- ditions of groundwater (Hölscher et al., 2010). In general, anism, greater extent of biodegradation is expected to provoke chlorinated benzenes are susceptible to oxidative transfor- greater changes in its isotopic composition. The magnitude of mation processes with decreasing chlorination degree but isotope fractionation during biodegradation of a selected con- the potential for anaerobic reductive dechlorination increases taminant is commonly determined in laboratory experiments with increasing number of chlorine substituents. Anaero- with pure or enrichment cultures. The results of these exper- bic reductive dechlorination of hexachlorobenzene (HCB) via iments are subsequently interpreted based on the Rayleigh pentachlorobenzene (PCB) and trichlorobenzenes (TCBs) to equation, that links degradation-induced shifts in isotope ra- dichlorobenzene (DCBs) has been reported in anoxic ground- tios (e.g. 13 C/12 C) to the degree of degradation (Elsner, 2010). waters and sediments (Capozzi et al., 2018; Kranzioch- The laboratory-derived isotope fractionation values can be af- Seipel et al., 2016; Sohn and Häggblom, 2016). terwards used to quantify the extent of biodegradation in situ, Under aerobic conditions, several bacteria are capable of provided that the environmental conditions and biochemi- growing on TCBs (with the exception of 1,3,5-TCB, which is cal reactions governing the degradation process in the labo- considered recalcitrant), the three isomers of DCB and MCB ratory experiments and at the field site are the same (Elsner, as the sole source of carbon and energy (Field and Sierra 2010). Alvarez, 2008). The degradative attack of chlorinated ben- Over the last decade, CSIA has been applied to study car- zenes by these aerobic bacteria is initiated with dioxyge- bon stable isotope fractionation during the reductive dechlo- nases to produce chlorocatechols by dihydrodiol dehydro- rination of various chlorinated compounds, and particularly genase. The catechols derived via this pathway are subse- chlorinated ethenes (Badin et al., 2014; Blázquez-Pallí et al., quently oxidized to chloromuconic acids, which are metab- 2019a; van Breukelen et al., 2017). However, the CSIA stud- olized further to intermediates of the Kreb’s cycle (Field and ies reported from pure cultures or field-derived microcosms Sierra Alvarez, 2008). The enhancement of in situ aero- with chlorinated benzenes have been more limited. No signif- bic bioremediation in aquifers can be achieved by direct icant changes in the carbon isotope composition have been injection of oxygen-release compounds (i.e. peroxides of observed during the dechlorination of highly chlorinated ben- calcium and magnesium), as used routinely for aerobic zenes (HCB, PCB, 1,2,3,5-tetrachlorobenzene) in microcosms biodegradation of hydrocarbons (Gholami et al., 2019). Un- derived from sediments (Sohn et al., 2018) whereas signifi- der anaerobic conditions, organohalide-respiring bacteria be- cant isotope fractionation during anaerobic reductive dechlo- longing to the genera Dehalococcoides and Dehalobacter can rination of 1,3-DCB and 1,4-DCB by methanogenic cultures use higher chlorinated benzenes as terminal electron ac- and 1,2,4-TCB by Dehalococcoides mccartyi strain CBDB1 have ceptors, transforming them to lesser chlorinated benzenes been reported (Griebler et al., 2004; Liang et al., 2014). The congeners (Adrian et al., 2000; Nelson et al., 2014). At the aerobic degradation of MCB and 1,2,4-TCB by groundwater heart of this process are membrane-bound reductive de- and sediments collected from a MCB-contaminated site and halogenases, which catalyze dechlorination of chlorinated a Pseudomonas sp. P51, respectively, was not accompanied by compounds to generate cellular energy. Anaerobic dechlo- a significant carbon isotope fractionation (Griebler et al., 2004; rination can be enhanced by direct injection of easily fer- Marchesi et al., 2018). Compared to chlorinated C1 - and C2 - mentable substrates (i.e. lactate, vegetable oils, molasses) into hydrocarbons, a lower bulk carbon isotopic fractionation dur- the groundwater that results in the generation of hydro- ing compound transformation is typically observed for chlo- gen (Blázquez-Pallí et al., 2019a). Although complete reduc- rinated benzenes due to the higher number of C atoms. For tive dechlorination of monochlorobenzene (MCB) via benzene the latter, the carbon isotopic fractionation at the reactive has been demonstrated in methanogenic sediments (Fung position(s) of the molecule is diluted by the relatively high et al., 2009; Liang et al., 2013), this dechlorination pathway number of C atoms located at the non-reactive ones. Carbon has been rarely observed in field sites under strong reducing isotope fractionation at non-reactive positions, so called sec- conditions. ondary isotope effect, is typically an order of magnitude lower 206 journal of environmental sciences 118 (2022) 204–213 (Elsner et al., 2005). In addition, the weak isotopic fractiona- (multi-parameter 3430 WTW, Weilheim, Germany). All param- tion of chlorinated benzenes has been also attributed to the eters were recorded once stabilized and the Eh values were re- masking effect associated to their low solubility (Sohn et al., ported according to the standard hydrogen electrode system 2018). (UH). Afterwards, samples for chemical characterization and In this study, we aimed at assessing the efficacy of a biore- microcosms construction were collected. For the establish- mediation treatment in the subsurface of a former chemi- ment of microcosms, groundwater with fine sediments was cal plant located at an industrial area of Barcelona where the collected from the bottom of the monitoring wells in amber production of predominantly pesticides caused a release of a autoclaved glass bottles, which were previously filled with N2 large volume of chlorinated compounds, mostly chlorinated gas to reduce bacterial contact with oxygen, and sealed with benzenes, into the groundwater. Consequently, we assessed septum-lined (PTFE) caps to minimise the adsorption of chlo- the redox conditions of the aquifer and the intrinsic biodegra- rinated compounds. Groundwater for chemical characteriza- dation potential of this site using field-derived microcosms tion was collected at 1 m above the bottom of the wells to avoid (established with groundwater from either the source zone or sediments. the contaminant plume) and testing for aerobic and anaero- The second groundwater sampling campaign was done in bic biostimulation. A second goal of this study was to explore February 2020 to collect samples from wells P1 and P2 for CSIA whether carbon isotope signatures of MCB and 1,4-DCB would experiments. Samples were collected as described above and allow monitoring of their biodegradation at this site. the hydrogeochemical parameters of groundwater were also measured in situ following the same procedure (Appendix A Table S1). 1. Materials and methods All groundwater samples were kept in the dark at 4º C until analysis. 1.1. Hydrogeological description of the aquifer 1.3. Laboratory microcosms The studied site is located in the province of Barcelona (Spain), at around 400 m NW of the Besòs river and near the coastline. Each microcosm consisted of 160 mL glass serum sterile bot- This site was significantly impacted by chlorinated aromatic tles sealed with Teflon-coated butyl rubber septa and alu- hydrocarbons contamination due to improper disposal prac- minium crimp caps and contained 115 mL of sampled ground- tices during its former use as pesticide manufacturing plant. water and fine sediments. The lithology at the site consists of Quaternary alluvial de- posits (deltaic facies with a maximum thickness of 50 m) over- 1.3.1. Assessment of the intrinsic biodegradation potential of laying Pliocene marls and clays. the site For the characterization of the contamination at the stud- Three treatments (aerobic, anaerobic and abiotic control) were ied area, five piezometers (S1, S2, S3, S4, S5) were installed prepared in parallel sets of microcosms constructed with in the source zone where the chemical manufacturing plant groundwater from the source zone (wells S1, S2 and S3). was operating and three piezometers (P1, P2, P3) were installed Each treatment consisted of four replicates. To investigate downgradient in the contaminant plume (Appendix A Fig. S1). whether the addition of lactate could enhance the anaerobic Groundwater samples were collected from the upper part (up reductive dichlorination of lower-chlorinated benzenes in the to 9.1 m below ground surface, Appendix A Table S1). The up- source zone, sodium lactate (3 mmol/L) was added into mi- per part is an unconfined alluvial aquifer mainly composed crocosms prepared in an anaerobic glovebox (anaerobic treat- of sand and gravel layers of elevated hydraulic conductivity. ment). The effect of oxygen on the aerobic degradation of The water table at the site is relatively shallow (from 2.8 to 5.5 lower-chlorinated benzenes was investigated in parallel mi- m below ground surface, Appendix A Table S1) and the main crocosms gassed with oxygen to overpressure (+1.4 bar) (aer- groundwater flow direction is towards the SW (Appendix A Fig. obic treatment). Abiotic controls consisted of microcosms in- S1). Previous chemical investigations performed in the period activated with nitric acid (HNO3 70%) to pH 2. Each microcosm 2009-19 in the source zone and the plume revealed the pres- was spiked with a mixture of lower-chlorinated benzenes pre- ence of several chlorinated compounds in groundwater (Ap- pared in ethanol to reach 50 μmol/L (nominal concentration) pendix A Tables S2 and S3). of MCB, 1,2-DCB, 1,3-DCB, 1,4-DCB, 1,2,3-TCB, 1,2,4-TCB, and 1,3,5-TCB. When either MCB or 1,4-DCB were supplied alone, 1.2. Collection of aquifer samples the stock solutions were prepared with acetone. To study whether oxygen could enhance the aerobic Two groundwater sampling campaigns were carried out. The biodegradation of MCB in the contaminant plume downgra- first one was done in November 2019 on monitoring wells dient from the source, two treatments (aerobic and abiotic from both the source zone (S1, S2, S3) and plume (P1, P2, P3) control) were prepared in parallel sets of microcosms con- using a submersible stainless-steel pump (diameter of 1.82 structed with groundwater from the plume (wells P1, P2 and inches and 12 V). After proper purging of the wells, hydro- P3). Each treatment consisted of four replicates. Aerobic mi- geochemical parameters (i.e., temperature, pH, electric con- crocosms were gassed with oxygen to overpressure (+1.4 bar) ductivity, redox potential (Eh) and dissolved oxygen (DO), and abiotic controls consisted of microcosms inactivated with Appendix A Table S1) were measured using a flowthrough acid (70% HNO3 ) to pH 2. Each microcosm was spiked with cell connected to the pump (to avoid contact with the at- 50 μmol/L (nominal concentration) of MCB unless otherwise mosphere) and equipped with the corresponding electrodes stated. journal of environmental sciences 118 (2022) 204–213 207 All microcosms were incubated at 25 ºC in the dark without Samples for 13 C-CSIA of MCB and 1,4-DCB were prepared at agitation. the laboratories of the research group on Stable Isotopes and Mineralogy (MAiMA) at the University of Barcelona (UB) and 1.3.2. Biodegradation experiments for CSIA analyzed with a GC (Agilent 6890) coupled to an IRMS (Ther- Groundwater samples collected from the plume (wells P1 and moFinnigan Delta Plus) at “Centres Científics i Tecnològics de P2) were purged with nitrogen for 20 min to remove chlo- la UB” (CCiT-UB). For each compound, aqueous samples were rinated solvents. Each experimental bottle consisted of 160 diluted to a similar concentration in 20 mL vials containing mL glass serum sterile bottles sealed with Teflon-coated butyl a 30 mm PTFE-coated stir bar. The vials were immediately rubber septa and aluminium crimp caps and contained 100 sealed with PTFE/Silicone septa and aluminum crimp caps. mL of a mixture of groundwater and fine sediments from wells Then, the sample solution was stirred at room temperature P1 and P2 (50 mL each). Two sets of experiments were estab- and the corresponding analyte was extracted during 20 min by lished with fourteen microcosms each containing 200 μmol/L headspace solid-phase micro extraction (SPME) using a man- MCB and 150 μmol/L 1,4-DCB (nominal concentration), respec- ual sampler holder equipped with a 75 μm Carboxen-PDMS tively. The concentration of MCB and 1,4-DCB was monitored fiber (Supelco, Bellefonte, PA) (Palau et al 2017). The GC was periodically through the experiment by headspace analysis as equipped with a Supelco SPB-624 column (60 m × 0.32 mm, described in section 1.4 and each microcosm was sacrificed at 1.8 μm film thickness; Bellefonte, PA). The injector was set at different extent of degradation by adding 300 μL HNO3 (70%) 250°C in split mode (1:10) and the oven temperature program through the septum. was kept at 60°C for 2 min, heated to 220°C at a rate of 8°C/min Two different controls were included: (i) controls killed and finally held at 220°C for 5 min. Helium was used as a car- with 300 μL HNO3 (70%) containing groundwater and sedi- rier gas with a gas flow rate of 1.8 mL/min. Isotopic working ments without addition of MCB and 1,4-DCB to check that both standards of MCB and 1,4-DCB were used to ensure accuracy of compounds were efficiently removed after purging with nitro- the isotopic measurements during the course of samples anal- gen, and (ii) controls containing sterilized distilled water with ysis and were measured interspersed in the same sequence. MCB and 1,4-DCB to account for abiotic transformations and The working standards had been referenced isotopically be- potential impurities from the stock solutions. All microcosms forehand towards the international standard Vienne Pee Dee were incubated at 25 ºC in the dark without agitation. Belemnite (VPDB) with an elemental analyzer (EA) coupled to an IRMS. Standards and samples were analyzed by duplicate 1.4. Analytical methods as a quality control and the precision (1σ ) of δ 13 C values based on the analysis of the standards was ≤ ± 0.3‰ for both com- The concentration of MCB, DCBs and TCBs in liquid samples pounds. from the microcosms was determined by static headspace gas chromatography. For this, 3-mL liquid samples from each mi- 1.5. Isotope data evaluation crocosm were withdrawn and filtered through a 0.22 μm filter (Millex-GV, Millipore) and subsequently transferred to a 10 mL Carbon isotope ratios of MCB and 1,4-DCB were measured at vial containing 3.5 mL saline solution (NaCl, 1 mol/L), sealed natural abundance and were reported using the delta notation immediately with a Teflon-coated stopper. The vial was placed (Eq. (1)): in a headspace sampler Agilent 7964 (Agilent Technologies, 13  Palo Alto, CA) and heated to 85 ºC for 25 min. Subsequently, R C/12 C sample 13 δ Csample = −1 (1) a 1-mL headspace sample was injected automatically into a R(13 C/12 C )standard gas chromatograph (GC, Agilent 6890 N) equipped with a flame where R is the isotope ratio (13 C/12 C). The relationship be- ionization detector (FID) operated at 300 ºC, and a column DB- tween isotope fractionation and the extent of MCB and 1,4- 624 (30 m × 0.32 mm). The initial column temperature was DCB biodegradation in laboratory experiments was evaluated 40 ºC (held for 2 min); and then temperature slopes were ap- by a modified form of the Rayleigh distillation (Eq. (2)): plied as follows: slope 20 ºC/min (2 min); slope 10 ºC/min (2   min), slope 20 ºC/min (3 min), slope 1 ºC/min (4 min), slope Rt δ 13 Ct + 1 5 ºC/min (2 min), slope 10 ºC/min (3 min), slope 20 ºC/min (1 ln = ln 13 = εbulk · ln f (2) R0 δ C0 + 1 min), slope 40 ºC/min (2 min), final temperature: 260 ºC (main- tained 2 min). The injector was set at 250 ºC in split mode (1:2). where R0 is the initial isotope ratio at the beginning of the The carrier gas was helium at 40 mL/min. transformation process and Rt and f the isotope ratio and The concentrations of hydrogen, carbon dioxide, methane the compound remaining fraction (f = C/C0 ), respectively, at a and oxygen were measured on 0.1-mL headspace samples us- point in time t. The bulk (i.e., compound-average) isotope frac- ing an Agilent 7820A GC equipped with a thermal conductivity tionation values (εbulk ) were quantified by least squares linear detector as described elsewhere (Trueba-Santiso et al., 2020). regression of Eq. (2) without forcing the regression through the The concentration of anions (SO4 −2 , NO3 − , Cl− ) was mea- origin (Scott et al., 2004) and the uncertainty corresponds to sured on a Dionex ICS-2000 ion chromatography system the 95% confidence interval (C.I.) derived from the standard (Dionex Corp., Sunnyvale, CA, USA) equipped with an IonPac deviation of the regression slope. AS18 anion-exchange column. The column was operated at 30 For a given substrate, intrinsic kinetic isotopic effects ºC and a flow rate of 1 mL/min. The injection volume was 25 (KIEs) during compound transformation are position spe- μL. The potassium hydroxide concentration of the eluent var- cific whereas ε bulk values are calculated from compound- ied from 25 mmol/L to 50 mmol/L along the 10 min analysis. average isotope data (Eq. (2)). Hence, observable ε bulk values 208 journal of environmental sciences 118 (2022) 204–213 must be converted into apparent KIEs (AKIEs) in order to ob- to be considered anaerobic in field assessments. Further- tain information about the underlying reaction mechanisms more, anaerobic oxidation of compounds such as benzene (Elsner et al., 2005). For the calculation and interpretation or 1,2-dichloroethane was demonstrated under iron (III)- and of AKIEs a hypothesis about the reaction mechanism, or as- nitrate-reducing conditions, respectively (Anderson et al., sumed reaction mechanism, is thus necessary. The effects of 1998, Dinglasan-Panlilio et al., 2006). Hence, additional infor- non-reacting positions within the molecule, as well as of in- mation is necessary for better characterization of chlorinated tramolecular competition, are then taken into account with benzenes biodegradation in the field. the Eqs. (3) and (4), respectively (Elsner et al., 2005), A clear downward trend of TCBs, DCBs, MCB towards lower concentration levels is observed between 2009-2019 in well n εrp ≈ ·ε (3) S1 and between 2015-2019 in well S2 (Appendix A Table x bulk S2). Despite this, the concentration of MCB, 1,4-DCB, 1,2,3- TCB and 1,2,4-TCB (only in well S2) in 2019 exceeded the 1 groundwater remediation goals established by the Catalan AKIEC =  εrp  (4) z· 1000 + 1 authorities (Appendix A Table S2). In contrast, the concen- tration of lower-chlorinated benzenes in well S4 and S5 in where ε rp is the isotopic fractionation at the reactive position, 2019 complied with the legal requirements (Appendix A Ta- “n” is the number of C atoms, “x” is the number of C atoms ble S2). In 2019, all monitoring wells with the exception of at reactive sites (i.e., atoms that would experience isotope ef- S5, P1 and P2 exceeded the groundwater remediation goals for fects in the given reaction) and “z” is the number of identical lindane, 2,4’-dichlorodiphenyldichloroethane (DDD), and 4,4’- reactive sites undergoing intramolecular competition. These DDD (Appendix A Tables S2 and S3). The concentration of lin- equations assume the absence of secondary isotope effects. dane decreased in the period 2015 to 2019 in wells S1, S2 and As mentioned above, this is a valid assumption for carbon be- S4 (Appendix A Table S2). Although the concentration of lin- cause secondary carbon isotope effects are usually not signif- dane was relatively low, anaerobic reductive dechlorination of icant compared to the primary ones (Elsner et al., 2005). lindane might slightly contribute to the generation of lower chlorinated benzenes (mainly 1,2-DCB, 1,3-DCB, MCB and ben- zene) in this anoxic aquifer (Zhang et al., 2020). 2. Results and discussion The analyses of chlorinated compounds across the site showed that in the wells located in the source (S1, S2 and 2.1. Site characterization S3), MCB, the three dichlorobenzene isomers (1,2-, 1,3-, and 1,4-DCB), and 1,2,3- and 1,2,4-TCB were detected at the high- The hydrogeochemical data collected from the wells during est concentration whereas the plume (P1, P2, P3) mainly con- the two sampling campaigns is summarized in Appendix A sisted of MCB and 1,4-DCB (Appendix A Tables S2 and S3). Table S1. In the monitored wells, the water table was located The observed spatial distribution of contaminants can be ex- at depths ranging between 2.8 and 5.5 m below ground sur- plained by the different physicochemical characteristics of the face. The pH and temperature values were on average 7.1 ± higher chlorinated benzenes and their dechlorinating prod- 0.1 and 20.3 ± 0.9 ºC, respectively, suitable for microbial activ- ucts. Under strong reducing conditions, organohalide respi- ity. Dissolved oxygen concentration ranged from 0.04 to 0.33 ration of higher chlorinated benzenes leads to lower chlori- mg/L, with exception of the value of 1.2 mg/L in well P1 (plume nated benzenes associated with an accumulation of MCB. The zone) in February 2020. This result suggests that groundwater low water solubility and strong retardation of higher chlori- at the site is mainly anoxic. The concentration of nitrate and nated compounds reduces their mobility in the by-passing sulphate was below the method detection limit (≤0.003 mg/L), water but favors the down gradient transport of DCBs and MCB which correlates well with the very low dissolved oxygen con- (Stelzer et al., 2009). centrations and the negative standard redox potentials mea- The isomer 1,2,4-TCB, which is a transient intermediate sured in wells S1, S3 and P1. The low EhSTD value (-39.7 mV) during the anaerobic reductive dechlorination of higher chlo- obtained in P1 in February 2020 suggest that the relatively high rinated benzenes, was the most abundant within TCBs in all DO concentration detected in this well (1.2 mg/L) might be due wells (<60% of total TCBs) (Appendix A Tables S2 and S3), to atmospheric oxygen contamination during the measure- suggesting that biodegradation was one of the natural atten- ment. A lower DO value of 0.33 mg/L was measured in P1 in uation processes for the higher chlorinated benzenes occur- November 2020. Positive standard redox potentials were found ring at the site (Tas et al., 2011). The dechlorination of singly in well S2, P2 and P3, which were in the typical range of iron flanked chlorine residue of 1,2,4-TCB producing 1,4-DCB is the (III)-reducing conditions (0-100 mV). thermodynamically more favorable reaction (Hölscher et al., Determination of the redox conditions can help to elu- 2010) and, accordingly, this DCB isomer accounts for more cidate the contaminant degradation pathways, which is es- than 60% of total DCB in wells P1, P2, P3 and S1. The prevalence sential information for (i) evaluating the natural attenua- of 1,4-DCB can be also explained by the fact that the most tion of chlorinated benzenes at contaminated sites and (ii) adjacent chlorine substituents are preferable used during predicting potential accumulation of toxic products. How- organohalide respiration of DCBs (Adrian and Gorisch, 2002). ever, aerobic biodegradation of chlorinated benzenes might Contaminated sites are usually impacted by mixtures of be sustained at very low DO concentrations. For instance, chlorinated benzenes, which complicates the evaluation of Gossett (2010) reported VC oxidation in transfer cultures specific DCBs biodegradation from analysis of degradation with DO concentrations below < 0.02 mg/L, low enough products because the same products can be formed from dif- journal of environmental sciences 118 (2022) 204–213 209 ferent precursors. For instance, under reducing conditions MCB can be formed from 1,4-DCB, 1,2-DCB or 1,2,4-TCB by anaerobic reductive dechlorination (Lawrence, 2006). In ad- dition, end products of the aerobic degradation pathways of chlorinated benzenes, i.e., inorganic carbon and Cl− , are ubiq- uitous and often occur at high background concentrations in groundwater. At the investigated site, elevated levels of chlo- ride were detected in all the wells (≥ 205 mg/L) but they can- not be directly related to dechlorination processes because the natural background concentration of chloride was not deter- mined in non-contaminated groundwater wells. Rather, this concentration of chloride is probably due to the intrinsic char- acteristics of this coastal aquifer. Therefore, investigation on additional tools, such as CSIA, for improved assessment of chlorinated benzenes biodegradation in contaminated field sites is warranted. 2.2. Aerobic and anaerobic microcosms established with groundwater from the source zone In the microcosms amended with oxygen, MCB and 1,4-DCB were degraded in all replicates regardless of the groundwater source. Abiotic controls and anaerobic treatments did not de- grade any of the investigated chlorobenzenes. Anaerobic mi- crocosms were monitored for 55 days, and methane was not produced during this period, indicating that strongly reduc- ing conditions were not achieved (data not shown). Absence of anaerobic biodegradation of lower chlorinated benzenes might be due to the inhibitory effect of co-contaminants in the medium. The aerobic microcosms established with groundwater from S1 showed a faster dechlorinating activity, with 71.4% of MCB degraded within 15 days and 81.1% of 1,4-DCB de- Fig. 1 – Degradation of MCB (Panel a) and 1,4-DCB (Panel b) graded within 45 days (Fig. 1a-b). At this degradation point, in aerobic microcosms containing groundwater from well MCB was reamended with 50 μmol/L, and 62% of MCB was S1. Red and green lines represent concentration of the degraded in 15 days, but the third dose of MCB, amended at respective chlorinated benzene in abiotic and aerobic a higher concentration (200 μmol/L), only started degrada- treatments, respectively. Blue and black lines represent tion after the addition of a mixture of sodium phosphate (20 oxygen and carbon dioxide concentration in aerobic mg/L), ammonium sulphate (50 mg/L) and oxygen (Fig. 1a). treatments, respectively (Panel c). Arrows indicate The second dose of 50 μmol/L 1,4-DCB, which was added at reamendment of the corresponding chlorinated benzene day 60, was degraded at 80.6% in 30 days (Fig. 1b). The standard (green), oxygen (white) and nitrogen/phosphorous source deviation became more significant in the later stages of the (black). Each curve shows the mean value of four replicates. degradation, likely due to the heterogeneity in the microbial Error bars indicate standard deviations of four replicates. ecology of the real groundwater used in these microcosms. A longer lag time and slower degradation rate of MCB and 1,4- DCB were observed in microcosms constructed with ground- waters from wells S2 and S3 (Appendix A Fig. S2). MCB and 1,4-DCB can be both utilized as sole source of carbon and en- ergy by a variety of bacteria, including the genera Pseudomonas, sources were mineralized in the microcosms, and therefore it Burkholderia, Xanthobacter, and Rhodococcus, among others, ini- cannot be assigned -at least, not exclusively- to MCB degrada- tiating their degradation with dioxygenases (Field and Sierra- tion (Appendix A Table S4). For instance, acetone and ethanol Alvarez, 2008). used as solvent in stock solutions were completely consumed Degradation of 1,4-DCB and MCB was concomitant with in all microcosms (data not shown), and other chlorinated a net increase of CO2 in the headspace (Fig. 1c and Ap- compounds and labile organic matter present in groundwa- pendix A Fig. S2). According to the stoichiometry of the oxi- ter but not monitored in this study would be susceptible to dation reaction, the amount of CO2 generated per each μmol aerobic degradation as well. of chlorinated benzene consumed would be 6 μmol (i.e., In parallel, four culture bottles per monitoring well were C6 H6 Clx + 7.5O2 → 6CO2 + 3H2 O + XCl− ). The observed ratio of amended with HCB crystals in anaerobic microcosms with CO2 produced against the CO2 expected from complete oxida- lactate to study whether lower-chlorinated benzenes were tion of chlorobenzenes degraded indicated that other carbon generated in this site from the anaerobic reductive dechlo- 210 journal of environmental sciences 118 (2022) 204–213 Appendix A Fig. S4). No MCB degradation was observed in abi- otic controls. Once demonstrated that MCB was degradable under aerobic conditions we aimed to study if DCBs, which were co-contaminants of MCB in these wells, might have an inhibitory effect on aerobic MCB degradation. For this, every bottle was reamended with MCB (200 μmol/L) and each of the four replicates constructed for each well were amended with either 1,2-DCB (50 μmol/L), 1,3-DCB (50 μmol/L) or 1,4- DCB (50 μmol/L). One of the replicates was not amended with DCB and remained as a control. The concentration of MCB re- mained unchanged until a nitrogen and phosphorus source was added (Fig. 2 and Appendix A Fig. S4). From that moment on, MCB and DCB started degradation and were rapidly con- sumed (Fig. 2, Appendix A Figs. S4, S5). These results show Fig. 2 – Degradation of MCB in aerobic microcosms that the three isomers of DCB and MCB were easily degrad- containing groundwater from well P2. Red and green lines able under aerobic conditions by the autochthonous bacteria represent concentration of MCB in abiotic and aerobic with no observable inhibitory effect. It also raised the ques- treatments, respectively. Blue and black lines represent tion whether the lack of biodegradation of 1,2- and 1,3-DCB in oxygen and carbon dioxide in aerobic treatments, microcosms from the source (section 2.2.) was due to the inhi- respectively. Arrows indicate reamendment of the bition provoked by some of the chlorobenzenes contained in corresponding chlorinated benzene (green), oxygen (white) the amended mixture. and nitrogen/phosphorous source (black). The pink arrow indicates that at this point the first replicate received 50 μmol/L 1,2-DCB, the second replicate 50 μmol/L 1,3-DCB, 2.4. Stable isotope fraction during aerobic degradation of and the third replicate 50 μmol/L 1,4-DCB. The fourth MCB and 1,4-DCB replicate did not receive DCB and acted as a control. Each The aerobic degradation of MCB and 1,4-DCB was accompa- curve shows the mean value of four replicates. Error bars nied by a slight enrichment of 13 C in their residual fraction. indicate standard deviations of four replicates. On the one hand, MCB was enriched in 13 C from -28.3‰ to - 27.2‰ when 86% was degraded (Fig. 3a). On the other hand, the initial isotope composition of 1,4-DCB was -29.1 ‰ and when rination of HCB. In 6 out of 12 bottles we detected 1,3,5-TCB 92% was degraded, the residual 1,4-DCB was enriched in 13 C to and 1,4-DCB as only byproducts after three months of incu- -26.7 ‰ (Fig. 3b). A resulting ε of -0.7‰ ± 0.1 ‰ and -1.0‰ ± 0.2 bation (see results of well S3 in Appendix A Fig. S3), indicat- ‰ was calculated for MCB and 1,4-DCB, respectively, by plot- ing that anaerobic reductive dechlorination of HCB is feasi- ting the experimental bottles results on a double-logarithmic ble under anoxic conditions in this contaminated site. These plot according to Eq. (2) (Fig. 3c-d). The good linear correlation results are in agreement with several studies showing pro- (R2 MCB = 0.96, R2 14-DCB =0.95) indicated that aerobic degrada- duction of DCB isomers and 1,3,5-TCB as final products of tion of MCB and 1,4-DCB are well described by the Rayleigh HCB by pure and mixed cultures under anaerobic conditions, model. with 1,3,5-TCB often as dominant end product (Adrian and The carbon isotopic fractionation (ε bulk ) during aerobic Görisch, 2002; Fennell et al., 2004; Wu et al., 2002). If this path- biodegradation of 1,4-DCB was not available in the literature way was active in the field, we would expect production of yet. This is a very valuable information for future studies since 1,4-DCB from 1,2,4-TCB in the previous anaerobic microcosms knowledge of the compound- and reaction-specific enrich- amended with lactate, because 1,2,4-TCB is an obligate inter- ment factor (ε bulk ) is required to assess the degree of biodegra- mediate in the production of 1,4-DCB via the HCB dechlorina- dation using CSIA. tion pathway (Fennell et al., 2004; Hölscher et al., 2010). A feasi- The EPA establishes as a criterion, based on the typical un- ble explanation is that some of the chlorinated benzenes con- certainty of ± 0.5 ‰ for δ 13 C measurements, that the values of tained in the mixture, which were absent in the microcosms δ 13 C in the compounds of interest must be enriched (less neg- amended with HCB, might exert an inhibitory impact on the ative) by 2 ‰ compared to values of δ 13 C in the undegraded organohalide respiration of these compounds (i.e., 1,2,4-TCB). compounds to demonstrate that significant fractionation oc- Self-inhibition by degradation products has been widely re- curs and ensure reliable interpretation of biodegradation (U. S. ported in the environment and, in the case of chlorinated ben- EPA, 2008). According to the determined ε bulk values for aero- zenes, 1,3-DCB, 1,4-DCB and 1,3,5-TCB are known to inhibit bic biodegradation, a shift of δ 13 C above 2 ‰ (i.e., ࢞δ 13 C > 2.0 HCB dechlorination (Chau et al., 2018). ‰) would be obtained after a degradation extent of approx- imately 94% and 86% of MCB and 1,4-DCB, respectively. This 2.3. Aerobic microcosms established with groundwater result indicates that qualitative and quantitative assessments from the plume of MCB and 1,4-DCB transformation by aerobic biodegradation would only be possible in the aquifer parts with very large The microcosms established with groundwater from wells P1, contaminant degradation. Previous studies showed negligible P2 and P3 showed a similar degradation pattern, consuming enrichment of 13 C associated to aerobic degradation of MCB two consecutive doses of 50 μmol/L MCB in 26 days (Fig. 2 and (Kaschl et al., 2009; Marchesi et al., 2018). However, the assess- journal of environmental sciences 118 (2022) 204–213 211 Fig. 3 – Percentage of degradation and carbon isotopic composition of MCB and 1,4-DCB in aerobic microcosms containing a mixture of groundwater collected from wells P1 and P2 (Panel a and b) and double logarithmic plot according to the Rayleigh equation of the carbon isotope ratio versus the residual concentration of MCB and 1,4-DCB (Panel c and d). The error bars show the one standard deviation (± 1σ) for duplicate measurements and doted lines represent the 95% confidence intervals of the linear regression. ment of aerobic degradation of MCB might be monitored by chlorine isotope fractionation in the light of a recent study highlighting enrichments of δ 37 Cl higher than 2.5 ‰ from aer- obic microcosms established with groundwater contaminated with MCB (Marchesi et al., 2018). The weak carbon isotope fractionation during aerobic biodegradation of MCB contrasts with the significant 13 C en- richment factor of -5.0 ‰ observed during anaerobic reduc- tive dechlorination of MCB (Liang et al., 2011). Accordingly, some studies have correlated the occurrence of anaerobic re- ductive dechlorination of MCB to benzene at field sites with a progressive enrichment of the δ 13 CMCB (Alberti et al., 2017; Fig. 4 – Aerobic biodegradation pathways and initial Kaschl et al., 2009; Stelzer et al., 2009). Similarly, for 1,4-DCB enzymatic dioxygenase reaction for (a) monochlorobenzene a much higher ε bulk value (-6.3‰ ± 0.2 ‰) compared to the and (b) 1,4-dichlorobenzene. one determined in this study (-1.0‰ ± 0.2 ‰) was determined during anaerobic reductive dechlorination in methanogenic laboratory microcosms (Liang et al., 2014). Although the weak carbon isotope fractionation in the aerobic pathway would DCB, the initial attack on the aromatic ring takes place via only allow tracing of in situ degradation in aquifer parts with the formation of two C-O bonds, leading to the formation of high extent of degradation, relatively high carbon isotope 3-chloro-cis-1,2-dihydroxycyclohexa-3,5-diene (Werlen et al., shifts (i.e., ࢞δ 13 C > 4.0 ‰) of MCB or 1,4-DCB in contaminated 1996) and 3,6-dichloro-cis-1,2-dihydroxycyclohexa-3,5-diene groundwaters would indicate that their degradation is con- (Spiess et al., 1995), respectively. For MCB, an AKIE of 1.0021 trolled by anaerobic reductive dechlorination. ± 0.0003 was calculated using n = 6, x = 4 and z = 2 in Eqs. In order to characterize the isotope effect at the reactive (3) and (4), assuming that the two carbon atoms involved dur- position of MCB and 1,4-DCB, apparent kinetic isotopic ef- ing deoxygenation of the aromatic ring are in ortho- and meta- fect values (AKIE) were calculated. In the presence of oxy- position. The AKIE of 1,4-DCB was calculated in the same way gen, chlorinated benzenes are generally transformed by a ring and a value of 1.0030 ± 0.0006 was obtained. The AKIEs of dioxygenase (stereospecific ring deoxygenation) that intro- MCB and 1,2-DCB are similar and very small, reflecting the ab- duces two OH-groups in neighboring positions in the aromatic sence of bond breakage during the first (rate-limiting) reaction ring (van der Meer et al., 1991) (Fig. 4). For both MCB and 1,4- step. Liang et al. (2011) suggested that the weak carbon iso- 212 journal of environmental sciences 118 (2022) 204–213 topic effect during aerobic degradation of MCB is associated Alberti, L., Marchesi, M., Trefiletti, P., Aravena, R., 2017. to the initial enzymatic reaction catalyzed by dioxygenases Compound-specific isotope analysis (CSIA) application for (rate-limiting C-O bond formation) whereas the larger carbon source apportionment and natural attenuation Assessment of chlorinated benzenes. Water. 9, 872 DOI:10.3390/w9110872. isotope effect observed for anaerobic degradation is likely the Anderson, R.T., Rooney-Varga, J.N., Gaw, C.V., Lovely, D.R., 1998. result from rate-limiting C-Cl bond cleavage. Anaerobic benzene oxidation in the Fe(III) reduction zone of The AKIE of MCB (1.0021 ± 0.0003) is slightly higher than petroleum-contaminated aquifers. Environ. Sci. Technol. 32 those obtained by Kaschl et al. (2009), from 1.0003 ± 0.0003 (9), 1222–1229. doi:10.1021/es9704949. to 1.0012 ± 0.0003. These values were calculated in this study ATSDR (Agency for Toxic Substances and Disease Registry), 2019. using the same parameters (n, x and z) indicated above. The Priority list of hazardous substances. Available: AKIEs of MCB and 1,4-DCB obtained here are also consistent http://www.atsdr.cdc.gov/SPL/index. (accessed on 09.07.2021). Badin, A., Buttet, G., Maillard, J., Holliger, C., Hunkeler, D., 2014. with those obtained for ring dioxygenation reaction of ben- Multiple dual C-Cl isotope patterns associated with reductive zene (from 1.002 ± 0.0003 to 1.004 ± 0.0006, Fischer et al., 2008), dechlorination of tetrachloroethene. Environ. Sci. Technol. 48, toluene (1.001 ± 0.001 to 1.006 ± 0.001, Vogt et al., 2008) and 9179–9186. doi:10.1021/es500822d. ethylbenzene (1.002, Dorer et al., 2014). Blázquez-Pallí, Rosell, M., Varias, J., Bosch, M., Soler, A., Vicent, T., et al., 2019a. Integrative isotopic and molecular approach for the diagnosis and implementation of an efficient in-situ 3. Conclusions enhanced biological reductive dechlorination of chlorinated ethenes. Water Res 167, 115106. doi:10.1016/j.watres.2019.115106. The laboratory microcosms established in this study were a Blázquez-Pallí, N., Shouakar-Stash, O., Palau, J., valuable tool to assess biodegradation strategies to remediate Trueba-Santiso, A., Varias, J., Bosch, M., et al., 2019b. Use of an aquifer contaminated with multiple chlorinated aromatic dual element isotope analysis and microcosm studies to hydrocarbons. Our laboratory tests provided a positive indica- determine the origin and potential anaerobic biodegradation tion that enhanced in situ aerobic bioremediation is a feasible of dichloromethane in two multi-contaminated aquifers. Sci. strategy to degrade MCB and DCBs in this site. Notwithstand- Total Environ. 696, 134066. doi:10.1016/j.scitotenv.2019. 134066. ing, there is a need for additional laboratory studies to as- Capozzi, S.L., Rodenburg, L.A., Krumins, V., Fennell, D.E., sess the potential inhibitory effect of some chlorinated com- Mack, E.E., 2018. Using positive matrix factorization to pounds on the biodegradability of low chlorinated benzenes in investigate microbial dehalogenation of chlorinated benzenes the source area. The enrichment factor for 1,4-DCB offers an in groundwater at a historically contaminated site. important contribution to the use of 13 C-CSIA in future stud- Chemosphere 211, 515–523. ies and it could be used to assess aerobic biodegradation of doi:10.1016/j.chemosphere.2018.07.180. 1,4-DCA in aquifer parts with high extent of biodegradation. Chau, A.T.T., Lee, M., Adrian, L., Manefield, M.J., 2018. Syntrophic partners enhance growth and respiratory dehalogenation of In addition, on the basis of previous studies, the occurrence of hexachlorobenzene by Dehalococcoides mccartyi strain significant changes in carbon isotope ratios of MCB and 1,4- CBDB1. Front. Microbiol. 9, 1927. doi:10.3389/fmicb.2018.01927. DCB in the field (i.e., ࢞δ 13 C > 4.0 ‰) might suggest the pre- Dinglasan-Panlilio, M.J., Dworatzek, S., Mabury, S., Edwards, E., dominance of anaerobic reductive dechlorination pathways. 2006. Microbial oxidation of 1,2-dichloroethane under anoxic conditions with nitrate as electron acceptor in mixed and pure cultures. FEMS. Microbiol. Ecol. 56 (3), 355–364. Acknowledgments doi:10.1111/j.1574-6941.2006.00077.x. Dorer, C, Vogt, C., Kleinsteuber, S., Stams, A.J.M., Richnow, H.H., 2014. Compound-specific isotope analysis as a tool to This work was supported by the Catalan Water Agency (No. characterize biodegradation of ethylbenzene. Environ. Sci. CTN1900901). Jordi Palau was also supported by the projects Technol. 48, 9122–9132. doi:10.1021/es500282t. CGL2017–82331-R (Spanish Ministry of Economy and Compet- Elsner, M., Zwank, L., Hunkeler, D., Schwarzenbach, R.P., 2005. A itiveness) and 2017SGR 1733 (Catalan Government). We thank new concept linking observable stable isotope fractionation to Emilio Orejudo and Joan Sanchez (Catalan Agency of Water) transformation pathways of organic pollutants. Environ. Sci. for their help on field work. Technol. 39, 6896–6916. doi:10.1021/es0504587. Elsner, M., 2010. Stable isotope fractionation to investigate natural transformation mechanisms of organic contaminants: principles, prospects and limitations. J. Environ. Monit. 12, Appendix A Supplementary data 2005–2031. doi:10.1039/c0em00277a. Fennell, D.E., Nijenhuis, I., Wilson, S.F., Zinder, S.H., Supplementary material associated with this article can be Häggblom, M.M., 2004. 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