Application of γ-PGA as the primary carbon source to bioremediate a TCE-polluted aquifer: A pilot-scale study

Chemosphere 237 (2019) 124449

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Application of g-PGA as the primary carbon source to bioremediate a TCE-polluted aquifer: A pilot-scale study S.G. Luo a, S.C. Chen b, *, W.Z. Cao c, W.H. Lin a, Y.T. Sheu a, C.M. Kao a, ** a

Institute of Environmental Engineering, National Sun Yat-Sen University, Kaohsiung City, Taiwan Department of Life Sciences, National Central University, Chung-Li, Taiwan c College of the Environment and Ecology, Xiamen University, Xiamen, China b

h i g h l i g h t s

g r a p h i c a l a b s t r a c t


g-PGA can provide substrate and NHþ 4 continuously and enhance TCE dechlorination.
Release of amine from g-PGA could produce ammonia for acidification control.
g-PGA has a small diameter, which prevents soil clogging after injection.
g-PGA causes variations in microbial diversity and dominant bacterial species.

a r t i c l e i n f o

a b s t r a c t

Article history: Received 7 February 2019 Received in revised form 13 July 2019 Accepted 23 July 2019 Available online 25 July 2019

The effectiveness of using gamma poly-glutamic acid (g-PGA) as the primary carbon and nitrogen sources to bioremediate trichloroethene (TCE)-contaminated groundwater was studied in this pilot-scale study. g-PGA (40 L) solution was injected into the aquifer via the injection well (IW) for substrate supplement. Groundwater samples were collected from monitor wells and IW and analyzed for TCE and its byproducts, geochemical indicators, dechlorinating bacteria, and microbial diversity periodically. Injected g-PGA resulted in an increase in total organic carbon (TOC) (up to 9820 mg/L in IW), and the TOC biodegradation caused the formation of anaerobic conditions. Increased ammonia concentration (because of amine release from g-PGA) resulted in the neutral condition in groundwater, which benefited the growth of Dehalococcoides. The negative zeta potential and micro-scale diameter of g-PGA allowed its globule to distribute evenly within soil pores. Up to 93% of TCE removal was observed (TCE dropped from 0.14 to 0.01 mg/L) after 59 days of g-PGA injection, and TCE dechlorination byproducts were also biodegraded subsequently. Next generation sequence (NGS) analyses were applied to determine the dominant bacterial communities. g-PGA supplement developed reductive dechlorinating conditions and caused variations in microbial diversity and dominant bacterial species. The dominant four groups of bacterial communities including dechlorinating bacteria, vinyl chloride degrading bacteria, hydrogen producing bacteria, and carbon biodegrading bacteria. © 2019 Elsevier Ltd. All rights reserved.

Handling Editor: Tsair-Fuh Keywords: Bioremediation Dechlorination Dehalococcoides Poly-g-glutamic acid Trichloroethylene

1. Introduction * Corresponding author. ** Corresponding author. E-mail addresses: [email protected] (S.C. Chen), [email protected]. tw (C.M. Kao). https://doi.org/10.1016/j.chemosphere.2019.124449 0045-6535/© 2019 Elsevier Ltd. All rights reserved.

Trichloroethylene (TCE) is one of the widely used cleaning agents, which usually results in the contamination of subsurface

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environments worldwide due to accidental spills or intentional discharges (Li et al., 2015). If significant amount of TCE is released into the subsurface, TCE could form a dense, non-aqueous-phase liquid (DNAPL) (Peng et al., 2016), which has a slow release and dissolution characteristic causing a long-term threat to the environment (Aydin-Sarikurt et al., 2016; Chunming et al., 2017). The mass transferred rate from the DNAPL to the subsurface environment is controlled many complex factors including soil properties, groundwater flow, and mass transport mechanisms, and thus, it becomes a lengthy and costly processes to cleanup DNAPL sites (Steffan and Schaefer, 2016). Because TCE is a volatile organic compound and it has a density higher than water, at many TCE spill sites, the residual DNAPL of TCE would persist within pore spaces or fractures. Thus, most of the TCE spills result in groundwater contamination, and TCE concentrations in the unsaturated zone usually become non-detectable. TCE site remediation must focus on the cleanup of dissolved phase contamination. Moreover, more economic approaches are desirable for groundwater remediation to provide a long-term control of contaminated groundwater (Chunming et al., 2017; Zhang et al., 2017a). Bioremediation is an environmental friendly and cost-effective technology that has been applied frequently for the cleanup of contaminated sites including TCE-polluted groundwater sites (Megharaj and Naidu, 2017). In situ bioremediation system applies the polluted aquifer as a treatment system for pollutants biodegradation (Ikehata et al., 2016; Zhang et al., 2017b). Some bacteria perform dehalorespiration processes, which use chlorinated organic compounds as electron acceptors during the remediation. This kind of in situ bioremediation via activation of dehalorespiration by exogenous electron donors is believed to be applicable methods to remediate sites contaminated by chloroethenes (Yohda et al., 2017). The feasibility of bioremediation is dependent on the presence of native microbial community of dechlorinating bacteria in the polluted sites (Kahlon, 2016). Factors that affect the bioremediation efficiency include microbial ecology (e.g., bacterial species, microbial diversity, bacterial population, enzyme activities), environmental conditions [e.g., pH, dissolved oxygen (DO), temperature], electron acceptors, electron donors, soil formation, nutrients, contaminant characteristics. Thus, conditions that help microbial growth and activity in environments will significantly enhance metabolic degradation of pollutants (Joshi, 2018; Sudheesh, 2018). Primary carbon source (primary substrate or electron donor) supplement is a potential approach to enhance the efficiency of in situ bioremediation of TCE
rez-de-Mora et al., 2018). polluted sites (Liu et al., 2017; Pe Reductive dechlorination mediated by dechlorinating bacteria plays a key role in the bioremediation of the chlorinated ethenes including TCE. Under reductive dechlorinating conditions, TCE can be dechlorinated to non-toxic end product (ethene or ethane) by replacing chlorine atoms with hydrogen (Wen et al., 2017), and the produced intermediates include dichloroethene isomers [cis-DCE, trans-DCE, 1,1-DCE (total DCE or t-DCE)] and vinyl chloride (VC) (Kanitkar et al., 2016; Lien et al., 2016). The following bacteria are capable of dechlorinating TCE and other chlorinated ethenes: Desulfomonile, Desulfovibrio, Geobacter, Desulfuromonas, Desulfitobacterium, and Dehalococcoides (DHC) (Kanitkar et al., 2016; Wen et al., 2017). For example, DHC can completely reduce TCE via DCEs and VC to ethene (Wen et al., 2017). Using DHC cultures to perform reductive dechlorination of chlorinated ethenes can be enhanced by supplying electron donor (carbon source or hydrogen) and maintaining optimal pH condition (around neutral condition) (Wen et al., 2017). Decreased pH due to the formation of organic acid during fermentation and release of HCl during dechlorination would result in the growth inhibition of dechlorinating bacteria and cause negative impact on reductive

dechlorination (Yang et al., 2017b). Primary carbon sources can be applied to provide electron donors and energy sources to enhance the reductive dechlorination (Wu et al., 2016; Tillotson and Borden, 2017). Effective substrate distribution in the soil pores has become a great challenge during the substrate injection process (Yang et al., 2018a). Poly-g-glutamic acid (g-PGA) [(NHCH(CO2H)CH2CH2CO)n] is a naturally occurring and anionic extracellular polymer, and it is composed of viscous homo-polyamide of D and L-glutamic acid units (Zhang et al., 2018). It is biodegradable and it is usually synthesized by Bacillus species (Sirisansaneeyakul et al., 2017). g-PGA has a potential to be used in different industries (healthcare, food, environment, pharmaceutical) (Zhang et al., 2017a; Bhat et al., 2018). g-PGA provides the ability for metal removal due to its cation exchange capacity (Clarke et al., 2017; Qiu et al., 2017b; Sheu et al., 2018; Yang et al., 2018b). Next generation sequencing (NGS) provides information about known species in the sample (Pelissari et al., 2017). With the right expert knowledge about the species, it can offer a prophyle of the microbiology (bioprophyle) (Geurkink et al., 2016). NGS technologies could characterize total microbial communities using the high
ndez throughput methods (Pelissari et al., 2017; Rodríguez-Ferna et al., 2018). Because g-PGA has biocompatible, environmentally acceptable, and biodegradable characteristics, it is hypothesized that it has a potential to be applied as a primary carbon source during the reductive dechlorination process of chlorinated ethenes. The primary objectives of this study included: (1) analyses of basic physical, chemical, and biological characteristics of g-PGA, (2) conduction of an in situ field-scale operation to evaluate the efficiency of using g-PGA as the carbon source for the enhancement of TCE dechlorination, and (3) application of molecular biology techniques including real-time polymerase chain reaction (PCR) and NGS to assess the changes of dominant microorganisms, microbial community, and microbial diversity. The laboratory study was to investigate the characteristics of g-PGA and mechanisms the TCE removal. The field-scale study was conducted to assess the feasibility and effectiveness of g-PGA injection on TCE-contaminated groundwater remediation. 2. Materials and methods 2.1. Site description A TCE-spill site was used for this in situ field-scale study. The site soils were silty sands, and the subsurface water table was 3e4 m below land surface. The measured hydraulic conductivity was 0.0071 cm/s with a porosity of 0.33. The groundwater temperature is in the range from 16 to 27  C. Fig. 1 is the site map showing the TCE plume, groundwater flow, g-PGA injection well (IW), three downgradient monitor wells (DW1 to DW3), and one upgradient well (UW). 2.2. Characteristics of g-PGA

g-PGA (with initial pH of 6.3, density of 1.03 g/cm3, and molecular weight of 300,000 Da) was provided by Vedan Enterprise Corp. in Taiwan for in situ injection. g-PGA was produced via the bioconversion pathway of the fermentation process. The size distribution of g-PGA globule was obtained using a microscope (Biological Microscopes E400, Nikon Co., Japan). The basic characteristic and surface structure of g-PGA globule was determined by the ESEM/EDS (Environmental Scanning Electron Microscope/Energy Dispersive Spectrometer) (FEI Quanta 200, FEI, USA). g-PGA globule distribution was determined by a laser diffraction particle size

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Fig. 1. Site map showing the locations of g-PGA injection wells, monitor wells, and groundwater flow direction.

analyzer (Coulter LS 100, Beckman Coulter, USA), and the microfluidizer zeta potential analyzer (ZetaSizer-2000, Malvern, UK) was applied for the zeta potential determination. The viscosity was analyzed by Brookfield DV3T (USA) using the method described in Esfe et al. (2018). 2.3. Pilot-scale study In this study, one injection well (labeled as IW) was installed for

g-PGA supplement and groundwater sampling and analyses (Fig. 1). Three monitoring wells (labeled as DW1, DW2, and DW3) and one upgradient well (labeled as UW) were installed in the downgradient and upgradient areas, respectively, groundwater quality monitoring. The downgradient monitor wells DW1, DW2, and DW3 were 2, 3, and 4 m downgradient of IW, respectively. In the field-scale study, g-PGA solution was injected into the subsurface via the IW to enhance the TCE dechlorination. Two different g-PGA injection events were conducted during the operational period on days 0 and 59 (after groundwater sampling event). For each g-PGA injection event, 20 L of g-PGA (mixed with 200 L of groundwater) was supplied for substrate supplement. The total of 220 L of g-PGA solution was to allow the g-PGA solution to be distributed around the injection point evenly. 2.4. Groundwater sample analyses Groundwater samples were collected from IW, UW, and DWs and analyzed for geochemical parameters [pH, dissolved oxygen (DO), sulfate, nitrate, oxidation-reduction potential (ORP), ferrous iron (Fe(II), methane, and total organic carbon (TOC)], organic

contaminants (TCE and its degradation byproducts), DHC, and microbial diversity. EPA Method 601 was applied for the analysis of organic compound in accordance with using a Tekmer Purge-andTrap Model LSC 2000 with a Perkin-Elmer Model 9000 Auto System Gas Chromatograph (GC). Methane was analyzed on a Shimadzu GC-9A GC using headspace techniques. Anions were analyzed by in chromatography (Dionex, USA), and a Total Carbon Analyzer (Shimadzu, Japan) was used for TOC determination. DO was measured by an Orion (USA) DO meter (Model 840) and pH was analyzed by an Accumet 1003 pH/ORP meter (Fisher Scientific, USA). Fe(II) was determined using the Perkin-Elmer Plasma II Inductively Coupled Plasma-Argon Emission Spectrometer (ICPAES) (USA). The major analytical procedures are described in Standard Methods (APHA, 2005).

2.5. Real-time PCR for DHC and NGS anlayses The DNA Purification kit (GeneMark Co., Taiwan) was applied for bacterial DNAs extraction from each groundwater sample (4 L). The fragments of DHC was amplified using the primer described in He et al. (2003). The analytical procedures for PCR and real-time PCR for DHC analyses are described in Sheu et al. (2016). The 16S rRNA genes in V3eV4 regions of were amplified following the procedures described in Klindworth et al. (2013). The PCR reactions were performed using the Phusion® High-Fidelity PCR Master Mix (New England Biolabs, USA). The libraries generated with NEBNext® UltraTM DNA Library Prep Kit for Illumina and quantified via qubit and real-time PCR (Thermo Fisher Scientific, USA) were sequenced to generate 300 bp paired-end raw reads. According to the unique barcode, the paired-end reads were

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assigned and truncated by cutting off the primer sequence and barcode. FLASH (V1.2.7) was used to merge paired-end reads (Bokulich et al., 2013). The high-quality clean tags were obtained by quality filtering on the raw tags under specific filtering conditions (Bokulich et al., 2013) via the Qiime (V1.7.0) (Guo et al., 2016). The tags were compared with the reference database (Gold database, USA) to detect chimera sequences (Edgar, 2013). The resulted highquality sequences were grouped into operational taxonomic units (OTUs) with UPARSE (version 7.1) using the “cluster_otus” command (Edgar, 2013). For OTU cluster and species annotation, sequences analysis were performed by Uparse software (Uparse v7.0.1001, USA) (Edgar, 2013; Qiu et al., 2017a). OTUs were assigned at sequences 97% similarity threshold. For each representative sequence, Mothur software was performed against the SSUrRNA database of SILVA Database (Quast et al., 2012). OTUs abundance information was normalized by a standard of sequence number corresponding to the sample with the least sequences. The Alpha diversities (Shannon index and Chao 1) were calculated with QIIME (Version 1.7.0) and displayed with R software (Version 2.15.3). 3. Results and discussion 3.1. Physical and chemical properties of g-PGA Results show that the measured viscosity of g-PGA was 1.28 cp. The relatively lower viscosity, higher molecular weight, and higher solubility would allow the g-PGA serve as a potential candidate for primary substrate supplement without causing clogging of soil pores. Fig. 2 presents the results of ESEM-EDS image of g-PGA. Fig. 3 is the zeta potential and particle size distribution of g-PGA. g-PGA globule had a diameter around 1.9 mm and a zeta potential of 96 mv. The micro-scale range of globule size and a negative charged state would enable g-PGA to disperse into soil pores and also transport to downgradient area. This could result in a larger radius of influence.

3.2. Influence of g-PGA injection on groundwater quality Table 1 presents the results of groundwater quality indicators in IW and four monitor wells on day 0 (before g-PGA injection) and day 7 (after g-PGA injection). The upgradient monitor well (UW) was located at the upgradient location of IW, and thus, influence of the groundwater quality due to g-PGA injection was not observed. Fig. 4a presents the changes of groundwater pH in IW and monitor wells due to g-PGA injection. The pH values were around 6.7 to 6.9 in site groundwater before g-PGA injection. However, pH in IW increased from 6.7 (before g-PGA injection) to 7.5 and 7.1 after 7 and 37 days of g-PGA injection, respectively. Results show that three downgradient monitor wells also had similar pH variation trends. No significant pH variation was observed in UW. The increased pH value in IW and downgradient monitor wells could be because of the release of amine (a functional group of g-PGA) in groundwater, which reacted with water and produced ammonia causing the alkaline condition. Therefore, pH drops because of fatty acids production after substrate fermentation processes occurred in other studies were not observed in this field-scale study (Sheu et al., 2016; Wen et al., 2017). Fig. 4b shows the changes of ammonia concentrations in IW and monitor wells. Increase in ammonia concentrations was detected because of the amine release from g-PGA. Results show that the ammonia concentrations in IW increased from 1.3 mg/L before gPGA to 216 mg/L after 18 days of g-PGA supplement. Ammonia concentrations slowly decreased to below 13 mg/L after 59 days of injection. Increase in ammonia concentrations were observed in IW and monitor wells after the second g-PGA addition on day 59. This confirmed that g-PGA addition caused the increased ammonia concentrations in site groundwater. The dissolved ammonia migrated with groundwater flow to farther downgradient area after 59 days of injection. No significant increase in ammonia concentrations was detected in UW. The increased ammonia concentration also reveals that g-PGA could be applied as the nitrogen source for

Fig. 2. ESEM-EDS image of g-PGA.

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Fig. 3. (a) Particle size distribution and (b) zeta potential distribution of g-PGA.

Table 1 Results of groundwater quality indicators in IW and monitor wells on day 0 (before g-PGA injection) and day 7 (after g-PGA injection). Well

UW

Day Day IW Day Day DW1 Day Day DW2 Day Day DW3 Day Day

0 7 0 7 0 7 0 7 0 7

pH sulfate nitrate ammonia Fe2þ

ORP

DO

mv

mg/L e

mg/L

mg/L

mg/L

mg/L mg/L

159 110 34 192 133 175 24 142 124 55

1.1 1.2 1.1 0.1 0.4 0.3 0.4 0.5 0.7 0.6

74.3 80.4 22.3 2.1 51.6 9.2 78.3 12.7 64.9 17.1

5.3 3.2 2.7 <0.1 3.2 <0.1 2.1 <0.1 2.6 0.2

1.1 2.3 1.3 195 1.6 28 1.5 21 1.2 16

0.20 0.89 0.52 16.4 0.41 5.7 0.28 4.3 0.02 2.6

6.9 6.8 6.8 7.5 6.7 7.3 6.7 7.2 6.9 7.2

methane

<0.01 <0.01 <0.01 2.45 <0.01 2.2 <0.01 1.78 <0.01 0.52

the growth of indigenous bacteria. The decreased ammonia in IW was due to the groundwater dilution and transport mechanisms, which caused the significant drop of ammonia concentrations in IW. Results show that groundwater pH remained neutral without significant change while TCE was degraded via the reductive

dechlorination. Results suggest that g-PGA could provide the effective buffering capacity in the subsurface. Thus, no pH control or buffer agent addition is needed when g-PGA is applied as the primary carbon source during the enhanced anaerobic bioremediation processes of TCE. Fig. 4c presents the changes of TOC concentrations in IW and monitor wells after g-PGA supplement. Increase in TOC concentrations was detected in IW, and TOC concentrations went up from 5 mg/L (before g-PGA injection) to 9820 mg/L after 7 days of g-PGA injection and slowly declined to 18 mg/L after 59 days of injection. Increased TOC concentrations were detected in three downgradient monitor wells after approximately 7e18 days of g-PGA supplement. The TOC concentrations went up to 4662 mg/L and 36 mg/L on day 7 in DW1 and DW2, respectively. TOC concentrations dropped to 15 and 8 mg/L after 59 days of g-PGA injection in DW1 and DW2, respectively. Increased TOC concentrations in groundwater could provide carbon sources to activate and enhance the anaerobic dechlorination process. The TOC concentrations in UW remained low (<6 mg/ L) during the investigation period, indicating that the intrinsic carbon was not significant enough to enhance the natural TCE biodegradation processes under different processes.

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Fig. 4. (a) Variations in pH in IW and monitor wells during the investigation period; (b) Variations in ammonia concentrations in IW and monitor wells during the investigation period; (c) Variations in TOC concentrations in IW and monitor wells during the investigation period; (d) Variations in DO concentrations in IW and monitor wells during the investigation period; (e) Variations in TCE concentrations in IW and monitor wells during the investigation period; (f) Variations in t-DCE concentrations in IW and monitor wells during the investigation period.

Fig. 4d presents the changes of DO values in IW and monitor wells after g-PGA supplement. The supplied g-PGA resulted in the depletion of DO and in dropping of ORP in IW (Table 1). DO concentrations in IW dropped from 1.1 mg/L (before g-PGA injection) to 0.1 mg/L and ORP dropped from 34 mv before g-PGA injection to 192 mv) after 7 days of the g-PGA injection. Similar DO and ORP decrease trends were also observed in downgradient wells (Fig. 4d and Table 1). Fig. 4e presents the changes of TCE concentrations in IW and monitor wells. TCE concentrations decreased to 0.14 mg/L (before

g-PGA injection) to 0.08 and 0.011 mg/L after 7 and 59 days of gPGA injection, respectively. TCE concentrations in IW further dropped to 0.001 mg/L on day 78 after the second g-PGA injection on day 59. Enhanced deductive dechlorination could be the main cause of TCE after g-PGA injection. The injected g-PGA enhanced the biological activity of TCE dechlorination causing the significant TCE removal under neutral conditions. TCE removal was also detected in downgradient monitor wells indicating that the injected g-PGA solution could disperse and transport to the downgradient area and causing the occurrence of TCE dechlorination

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within the plume. Because the main TCE contaminants in DW2 and DW3 came from the upgradient area, the decrease in TCE concentrations in the upgradient areas (areas around IW and DW1) of DW2 and DW3 would also affect the detected TCE concentrations in DW2 and DW3. Results also imply that high TOC concentrations would shift the redox potential from aerobic conditionals to anaerobic conditions and enhance the anaerobic biodegradation process. Again, no obvious TCE removal was detected in UW, and this also confirmed that the injected g-PGA played a key role in TCE dechlorination and plume migration control. Fig. 4f presents the changes of t-DCE (total DCE isomers) concentrations in IW and monitor wells. Results in IW show that t-DCE concentrations increased from 0.01 on day 0e0.058 mg/L on day 18, and then dropped to 0.02 mg/L on day 58. After the second g-PGA injection on day 59, t-DCE concentrations further dropped to 0.018 mg/L on day 78. Similar trends of t-DCE variations were observed in the downgradient monitor wells. However, no significant change of the t-DCE concentrations was detected in UW. Injected g-PGA caused the TCE dechlorination, and the TCE degradation byproducts (t-DCE) were produced. The produced tDCE byproducts could be degraded without causing any accumulation when the primary substrates (g-PGA) was sufficient as low levels of VC and ethene [varied from below detection limit (0.001 mg/L) to 0.012 mg/L] were detected in IW and downgradient monitor wells (data not shown). This implies that t-DCE was further dechlorinated to less chlorinated byproduct (VC) and the end product (ethene). The results also indicate that TCE could be dechlorinated completely by indigenous dechlorinating bacteria when g-PGA was used as the substrate, and thus, accumulation of tDCE and VC was not observed. Results from our previous studies show that occurrence of VC accumulation might be observed in some TCE spilled sites due to the acidification problem caused by the production of fatty acids through the anaerobic reaction (Sheu et al., 2016). DHC growth was inhibited because of the pH drop, and thus, a complete TCE dechlorination could not be observed (Yang et al., 2017a; Dong et al., 2018; Saiyari et al., 2018). In this study, a neutral condition was maintained, and thus, VC accumulation could be prevented. Therefore, enhanced reductive dechlorination of TCE due to the addition of g-PGA could be concluded.

3.3. Quantification of DHC Fig. 5 presents the DHC population in groundwater collected

Fig. 5. Amounts of DHC genes in IW and monitor wells at different sampling events.

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from IW and monitor wells. DHC populations in IW, DW1, DW2, and DW3 increased from 1.03  103, 6.3  103, 3.05  103, and 2.12  103 gene copies/L to 7.2  106, 6.18  105, 2.2  106, and 1.4  106 gene copies/L, respectively, from days 0e37. In IW and three monitor wells, increased trend of DHC populations was observed. However, increased DHC population was not detected in UW during the investigation period. Members of DHC have been found that they could reductively dechlorinate TCE if primary
rez-de-Mora et al., substrates were provided (Ismaeil et al., 2017; Pe 2018). The increased DHC population in IW and downgradient monitor wells confirmed the effectiveness of g-PGA addition on the enhanced TCE dechlorination. 3.4. NGS analyses NGS was performed to assess the responses of microbial community patterns due to the g-PGA addition. Results showed that a total of 1583511 and 1310912 high quality reads that cluster OTUs was observed in UW and IW-59, respectively. Alpha diversity indices (Richness, Chao1, ACE, Shannon diversity, Simpson diversity, Inverse Simpson, Shannon evenness, Simpson evenness and Good's coverage) were calculated using mothur. In addition to Good's coverage, rarefaction curve (Supplementary Figs. 1 and 2) indicates that the microbial communities were well sampled. Table 2 shows the identified bacteria that were responsible to the dechlorination of chlorinated compounds or biodegradation of gPGA under different mechanisms and conditions. Fig. 6 presents the distribution of identified bacterial genus involved in dechlorination processes in groundwater samples collected from UW on day 0 and IW on days 0 and 59. Fig. 7 presents the distribution of identified bacterial genus involved in dechlorination processes in groundwater samples collected from UW on day 0 and DW3 on days 0 and 59. Fig. 8 shows that bacterial diversity distribution in groundwater samples collected from UW on day 0, IW on days 0 and 58, and DW3 on days 0 and 59. Results from Fig. 7 and Table 2 show that the presence of some bacteria related to TCE degradation. These bacteria are Anaerovorax (Liu et al., 2014), Bacillus (Lampis et al., 2014; Ozaki et al., 2017), Citrobacter (Ramprakash and Muthukumar, 2018), Clostridium (Ainala et al., 2016), Gracilibacter (Lee et al., 2006), Hydrogenophaga (Eamrat et al., 2017), Macellibacteroides (Huang et al., 2016) and Propionispora (Bengelsdorf et al., 2015). After adding g-PGA, the relative percentage of the amounts of Dehalogenimonas in microbial community increased to 0.79% on day 59, indicating that Dehalogenimonas could dechlorinate VC to ethene under reductive dechlorination processes in IW. The percentage sequence of dechlorinating bacteria such as increased from 0.09% on day 0e0.79% on day 59 indicating that the efficiency of TCE dechlorination could be significant enhanced after g-PGA addition (Baldwin et al., 2017; Edwards et al., 2018). Results (Fig. 7) also show that the percentage of hydrogen producing bacteria increased from 1.4% on day 0e28.34% on day 59, indicating that more hydrogen could be produced after g-PGA addition. The identified site bacteria including Acidovorax (Sheu et al., 2018), Clostridium (Ainala et al., 2016), Comamonas (Shimodaira et al., 2016), Hydrogenophaga (Eamrat et al., 2017), Pseudomonas (Singha et al., 2017), Rhodoferax (Lien et al., 2016) and Stenotrophomonas (Sheu et al., 2018) had the capabilities and potential to enhance TCE dechlorination. Geobacter, Rhodofera, and Sphingobium were able to dechlorinate chlorinated organic compounds Clostridium could synthesize vitamin B12 to enhance the growth of DHC (increased from 1.40% before g-PGA addition to 28.34% on day 59 after g-PGA addition) (Ahmad et al., 2017; Sheu et al., 2018). Bacillus (Lampis et al., 2014; Ozaki et al., 2017), Comamonas (Shimodaira et al., 2016), Curvibacter (Sheu et al., 2018)

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Table 2 Functions of the identified bacteria in soil samples. Bacteria

Function of identified bacteria

References

Acidaminobacter Acidovorax

NO 3 reduction after addition of lactate Chlorobenzenes degradation TCE degradation Phenol degradation Hydrocarbon degrading Fermentative metabolism Fermentative metabolism Reduction of selenite H2 production H2 production H2 production Alcohol fermentation Synthesize a B12 vitamin known as pseudocobalamin Cometabolization cis- and trans-dichloroethenes Iron reduction Polychlorinated biphenyl degradation Denitrification Capable of dihaloelimination to VC and ethene Sulfate reduction Perchloroethene degradation Sulfate reduction Iron reduction Chlorinated hydrocarbon dechlorination Fermentans Debitrification TCE degradation Fermentans Metabolization of monosaccharides and disaccharides Dinitrogen fixation Fermentation Dechlorination of chlorinated compounds TCE and VC degradation Biodegradation of polycyclic aromatic hydrocarbons TCE dechlorination Denitrification

Li et al. (2016) (Sheu et al., 2018) (Sheu et al., 2018) Ahmad et al. (2017) Dahal et al. (2017) Liu et al. (2014) Ozaki et al. (2017) Lampis et al. (2014) Ramprakash and Muthukumar (2018) Ainala et al. (2016) Rabemanolontsoa et al. (2017) Veeravalli et al. (2017) Helliwell et al. (2016) Shimodaira et al. (2016) (Sheu et al., 2018) (Sheu et al., 2018) (Sheu et al., 2018) Baldwin et al. (2017) Lien et al. (2016) Lien et al. (2016) Gam et al. (2018) Li et al. (2016) Lien et al. (2016) Lee et al. (2006) Eamrat et al. (2017) (Sheu et al., 2018) Rout et al. (2017) Huang et al. (2016) Hernandez et al. (2015) Bengelsdorf et al. (2015) (Sheu et al., 2018) Lien et al. (2016) Singha et al. (2017) (Sheu et al., 2018) (Sheu et al., 2018)

Acinetobacter Anaerovorax Bacillus Citrobacter Clostridium

Comamonas

Curvibacter Dehalogenimonas Desulfosporosinus Desulfovibrio Geobacter Gracilibacter Hydrogenophaga Macellibacteroides Methylobacter Propionispora Pseudomonas Rhodoferax Sphingobium Stenotrophomonas Thauera

Fig. 6. Distribution of identified bacterial species involved in dechlorination processes in groundwater samples collected from UW on day 0 and IW on days 0 and 59.

Desulfosporosinus (Lien et al., 2016), Desulfovibrio (Gam et al., 2018), Hydrogenophaga (Eamrat et al., 2017), Methylobacter (Hernandez et al., 2015) and Thauera (Sheu et al., 2018), which were identified in site groundwater could consume primary substrates under different redox conditions to make an environment favoring the reductive dechlorianting conditions (Desulfovibrio increased from 0% before g-PGA addition to 0.09% on day 59 after g-PGA addition).

In IW, results show that Acinetobacter, Anaerovorax, Bacillus, Clostridium, Desulfovibrio, Geobacter, and Hydrogenophaga were detected and increased microbial diversity was observed after gPGA addition on day 59 (Fig. 6 and Table 2). Among these identified bacteria, Hydrogenophaga increased from 0.14% before g-PGA addition to 6.01% on day 59. This indicates that the addition of gPGA produced an environment, which was suitable for the growth

S.G. Luo et al. / Chemosphere 237 (2019) 124449

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Fig. 7. Distribution of identified bacterial species involved in dechlorination processes in groundwater samples collected from UW on day 0 and DW3 on days 0 and 59.

Fig. 8. Bacterial diversity distribution in groundwater samples collected from UW on day 0, IW on days 0 and 59, and DW3 on days 0 and 59.

of Clostridium. Dehalogenimonas increased from 0.09% before g-PGA addition to 0.79% on day 59 after g-PGA addition and Desulfovibrio increased from 0.02% before g-PGA addition to 1.13% on day 59. Dehalogenimonas and Desulfovibrio could activate the sulfate reduction process using sulfate as the electron acceptor and g-PGA as the electron donor, and result in a decrease in sulfate and TOC concentrations (Lien et al., 2016; Matassa et al., 2016; Gam et al., 2018) Subsequently, this could favor the reductive dechlorination process (Matassa et al., 2016; Gam et al., 2018). In IW, the percentage sequence of dechlorinating bacteria increased from 2.06% on day 0e9.84% on day 59, indicating that the efficiency of TCE dechlorination could be significantly enhanced after g-PGA addition (Fig. 8). Results also show that the percentage sequence of hydrogen producing bacteria in DW3 increased from 1.4% on day 0e28.3% on day 59, indicating that more hydrogen could be produced after g-PGA addition. Thus, g-PGA addition could cause an obvious change of the microbial community.

4. Conclusions In this pilot-scale study, the effectiveness of g-PGA supplement on TCE dechlorination was evaluated. The following conclusions can be derived: 1. g-PGA resulted in the increase in TOC and drop of DO in site groundwater, which changed the subsurface environment from aerobic to anaerobic conditions. Thus, the reductive dechlorination of TCE was activated, which resulted in the decrease in TCE concentrations. 2. NGS analyses show that g-PGA supplement caused variations in microbial diversity and dominant bacterial species. Results from NGS study show that four groups of bacterial communities were identified in site groundwater: (1) dechlorinating bacteria, (2) VC degrading bacteria, (3) hydrogen producing bacteria, and (4) carbon (primary substrate) biodegrading bacteria. 3. Neutral pH could be maintained in site groundwater after g-PGA supplement. This was due to the release of amine functional

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groups from g-PGA, which reacted with water and form ammonia. 4. Results suggest that g-PGA could be applied as a potential primary carbon and nitrogen sources for the enhancement of in situ TCE dechlorination. Injection of g-PGA into TCEcontaminated groundwater would create a biological barrier around the injection zone and result in the TCE dechlorination inside the barrier. This could contain the TCE plume and prevent its further migration for downgradient receptor protection. Therefore, the environmental risks in the polluted or downgradient zones due to the TCE contamination can be decreased. 5. The future research needs for g-PGA application needs to be focused on g-PGA reagent modification to extend its life in the field remediation. For example, emulsified oil, biopolymer, and gelatin can be mixed with g-PGA to prepare a long-term slowreleasing site applicable substrate. Acknowledgements This study was funded by the research project supported by the Taiwan Environmental Protection Administration (EPA) (Contract No. 08RMAE001). The authors would like to thank the personnel at Taiwan EPA, Vedan Enterprise Corp., and Sun Dream Environmental Technology Corp. in Taiwan for their direction and help during the research period. Appendix A. Supplementary data Supplementary data to this article can be found online at https://doi.org/10.1016/j.chemosphere.2019.124449. References APHA, 2005. Standard Methods for the Examination of Water and Wastewater, twenty-second ed. American Public Health Association, APHA-AWWA-WEF, Washington, DC, USA. Ahmad, S.A., Shamaan, N.A., Syed, M.A., Khalid, A., Ab Rahman, N.A., Khalil, K.A., Dahalan, F.A., Shukor, M.Y., 2017. Meta-cleavage pathway of phenol degradation by Acinetobacter sp. strain AQ5NOL 1. Rendiconti Lincei 28, 1e9. Ainala, S.K., Seol, E., Kim, J.R., Park, S., 2016. Effect of culture medium on fermentative and CO-dependent H2 production activity in Citrobacter amalonaticus Y19. Int. J. Hydrogen Energy 41, 6734e6742. Aydin-Sarikurt, D., Dokou, Z., Copty, N.K., Karatzas, G.P., 2016. Experimental investigation and numerical modeling of enhanced DNAPL solubilization in saturated porous media. Water, Air, Soil Pollut. 227, 441. Baldwin, B.R., Taggart, D., Chai, Y., Wandor, D., Biernacki, A.,L., Sublette, K.,T., Wilson, J., Walecka-Hutchison, C., Coladonato, C., Goodwin, B., 2017. Bioremediation management reduces mass discharge at a chlorinated DNAPL site. Groundwater Monitoring & Remediation 37, 58e70. Bengelsdorf, F.R., Poehlein, A., Schiel-Bengelsdorf, B., Daniel, R., Dürre, P., 2015. Genome sequence of the acetogenic bacterium oxobacter pfennigii DSM 3222T. Genome Announc. 3, 01408e01415. Bhat, A.H., Rehman, W.U., Khan, I.U., Khan, I., Ahmad, S., Ayoub, M., Usmani, M.A., 2018. 15 - nanocomposite membrane for environmental remediation. In: Jawaid, M., Khan, M.M. (Eds.), Polymer-based Nanocomposites for Energy and Environmental Applications. Woodhead Publishing, pp. 407e440. Bokulich, N.A., Subramanian, S., Faith, J.J., Gevers, D., Gordon, J.I., Knight, R., Mills, D.A., Caporaso, J.G., 2013. Quality-filtering vastly improves diversity estimates from Illumina amplicon sequencing. Nat. Methods 10, 57. Chunming, S., Robert, W.P., Thomas, A.K., Mark, T.W., Suzanne, K.O.H., Jacqueline, W.Q., Nancy, E.R., 2017. Long-term performance evaluation of groundwater chlorinated solvents remediation using nanoscale emulsified zerovalent iron at a superfund site. In: Sung Hee, J. (Ed.), Applying Nanotechnology for Environmental Sustainability. IGI Global, Hershey, PA, USA, pp. 92e111. Clarke, D.E., Pashuck, E.T., Bertazzo, S., Weaver, J.V.M., Stevens, M.M., 2017. Selfhealing, self-assembled b-sheet peptideepoly(g-glutamic acid) hybrid hydrogels. J. Am. Chem. Soc. 139, 7250e7255. Dahal, R.H., Chaudhary, D.K., Kim, J., 2017. Acinetobacter halotolerans sp. nov., a novel halotolerant, alkalitolerant, and hydrocarbon degrading bacterium, isolated from soil. Arch. Microbiol. 199, 701e710. Dong, J., Dong, Y., Wen, C., Gao, S., Ren, L., Bao, Q., 2018. A 2D tank test on remediation of nitrobenzene-contaminated aquifer using in-situ reactive zone with emulsified nanoscale zero-valent iron. Chemosphere 206, 766e776. Eamrat, R., Tsutsumi, Y., Kamei, T., Khanichaidecha, W., Tanaka, Y., Kazama, F., 2017.

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