Chloroethene tetrachloroethene (PCE) and trichloroethene (TCE) are amongst the most abundant pollutants of groundwater and soil due to their frequent use in industrial applications. These pollutants can be biodegraded through natural or enhanced anaerobic reductive dechlorination, where chloroethenes serve as electron acceptors and molecular hydrogen and acetate, both released as by-products of organic substrate fermentation reactions, are used by the dechlorinating bacteria as electron donors and as carbon sources, respectively [1
]. During this process, PCE is converted stepwise to TCE by removing one chlorine atom and replacing it with a hydrogen atom; likewise, trichloroethene (TCE), is primarily converted to cis
-1,2-DCE, then to vinyl chloride (VC), and finally to ethene [2
Anaerobic reductive dechlorination of chloroethenes is restricted to just a few bacterial genera (hereafter collectively referred to as anaerobic dechlorinators). Those capable of sequentially dechlorinating PCE or TCE down to cis
-1,2-DCE include Dehalobacter
] and Desulfitobacterium
]. Dehalococcoides mccartyi
] and Dehalogenimonas
] are anaerobic dechlorinators known to gain energy through dechlorination of DCE to VC and eventually to ethene using the reductive dehalogenase enzymes BvcA and VcrA [14
] or similar ones in the case of Dehalogenimonas
spp. Despite the presence of anaerobic dechlorinators, cis
-1,2-DCE and VC often accumulate in groundwater as the sequential steps of the reductive dechlorination process are less and less favourable thermodynamically and kinetically [15
], and/or the conditions for complete dechlorination are not always optimal.
Under aerobic conditions, chloroethenes can be oxidised both cometabolically and metabolically. During cometabolic oxidation, chloroethenes are only degraded into non-toxic end-products fortuitously when degrading enzymes are produced for degradation of bacterial growth substrates such as methane, ethene, ammonium or aromatic pollutants. Cometabolic degradation has been shown for all chloroethenes, though only rarely described for PCE [16
]. Aerobic cometabolic oxidation is related to certain aerobic bacteria, such as ethene-oxidisers (etheneotrophs) and methane-oxidisers (methanotrophs) [17
]. Methanotrops employ soluble and particulate methane monooxygenases (sMMO and pMMO, respectively) to oxidise methane as a primary growth substrate. Both sMMO and pMMO are also capable of fortuitous oxidation of chloroethenes. The sMMO have a broader substrate range than pMMO, and are more efficient at degrading chlorinated ethenes [20
]. The gene mmoX
, which encodes the sMMO α subunit, and pmoA
, which encodes the pMMO α subunit, are used as biomarkers of chloroethene cometabolic potential in groundwater [20
Etheneotrophs can cometabolise VC and DCE when growing on ethene as a primary growth substrate, while several pure etheneotrophic strains, such as Mycobacterium
, can also utilise VC as their sole carbon and energy source [18
]. Etheneotrophs, when growing on ethene and VC, express a soluble alkene monooxygenase (AkMO), transforming VC to epoxide chlorooxirane, which is further metabolised to 2-chloro-2-hydroxyethyl-CoM by epoxyalkane:coenzyme M transferase (EaCoMT) [19
]. The genes etnC
encode the α subunit of AkMO and the EaCoMT, respectively, and serve as emerging biomarkers for ethenotroph-mediated aerobic biodegradation potential, though they do not distinguish between metabolic and cometabolic biodegradation pathways [20
Two degradation pathways have been proposed as regards aerobic metabolic degradation of cis
-1,2-DCE. Both of them involve degradation through monooxygenase-catalysed epoxidation [25
], with the initial step catalysed by cytochrome P450 monooxygenase. Epoxides can be degraded subsequently either by epoxyalkane, coenzyme M transferase or through formation of glutathione conjugates [26
Only a few studies have focused on parallel presence of anaerobic dechlorinators and aerobic methanotrophs or ethenotrophs at contaminated sites. Liang et al. [20
] studied the potential for VC degradation at six contaminated sites based on abundance and expression of VC biodegradation genes, and suggested that both ethenotrophs and anaerobic VC dechlorinators simultaneously contributed to VC biodegradation at the sites with high VC attenuation rates. Richards et al. [19
] investigated spatial relationships between functional genes of ethenotrophs, anaerobic VC dechlorinators and methanotrophs in aquifer soil samples collected at a contaminated site, and found that functional genes of all the three bacterial guilds coexisted in 48% of the samples that appeared to be anaerobic.
These results attracted our interest to further assess the potential of using the alternate anaerobic/aerobic biodegradation of chloroethenes as a practical remedial tool, which can eliminate frequent accumulation of cis-1,2-DCE and VC in groundwater. The goal of this study is to investigate this potential of ongoing activities of both anaerobic and aerobic chloroethene degraders at a large number of remediated sites affected by biostimulation to different levels and to estimate limiting conditions for these microbial degradation processes. For such scanning, the qPCR data and hydrogeochemical parameters were analysed by advanced statistical methods.
3. Results and Discussion
3.1. Results of Chemical Analyses
For a full summary of the chemical analyses, see the supplementary material
(Supplementary Table S2
). As most samples in this study were collected from aquifers affected by historical or on-going remediation using biostimulation via delivery of organic carbon, the laboratory analysis revealed mostly anoxic redox processes. Ongoing methanogenesis was detected based on the applied criteria (≥0.5 mg/L of methane) in 30 of 49 groundwater samples analysed; however, strict methanogenesis was only detected in four samples, the remaining 26 samples displaying criteria for more than one redox process were, mostly including Fe(III) reduction (24 samples). Fe(III) reduction alone, or in combination with Mn(IV) or NO3−
reduction, was only identified in nine samples.
With regard to concentrations of individual chloroethenes, the prevailing anoxic conditions, favourable for reductive dechlorination, resulted in sequential degradation of the parent contaminants (PCE and/or TCE) down to cis-1,2-DCE, VC and ethene. An average Cl no. of 1.5 indicated an advanced state of reductive dechlorination. In 11 of the 49 samples the Cl no. was even below 0.5, showing almost complete dechlorination by biostimulation of anaerobic biodegradation performed at these sites.
In all the collected samples, cis-1,2-DCE was the dominant DCE isomer. The ratio of trans-1,2-DCE to cis-1,2-DCE was below 2.2% in all the samples (mean ratio 0.39%) and concentrations of 1,1-DCE were similar to trans-1,2-DCE. Therefore, data for trans-1,2-DCE and 1,1-DCE were not included into the final data set for statistical analysis.
Of the 49 samples taken, 43 did not contain any other contaminants (in addition to chloroethenes) in significant concentrations (molar mass of the sum of co-contaminants below 2.5% of the molar mass of the sum of chloroethenes in the respective sample was used as the criterion). Six samples contained significant concentrations of co-occurring contaminants: at contaminated site #6 (two samples), the main contaminants were chloroform and 1,2-dichloroethane, whereas toluene was the dominant contaminant in the samples collected at site #12, although, historically, the site was dominantly contaminated by chloroethenes. Although co-occurring contaminants present in groundwater might affect both anaerobic reductive dechlorination and aerobic biodegradation, they were not included in the final data set for statistical analysis as their occurrence was limited and scattered.
Acetylene as an intermediate of abiotic β-elimination of chloroethenes [41
] was detected in seven of the 49 samples taken. It was present at sites #2, #3 and #5 where zerovalent iron materials were injected together with the whey in the past. It can be concluded that abiotic β-elimination has contributed to the degradation of chloroethenes at these sites.
3.2. Results of qPCR
qPCR revealed the frequent occurrence of both aerobic and reductive biomarkers. Presence of the ethenotroph functional genes etnC
was confirmed in 44 of 49 (90%) samples analysed, as were the methanotroph functional genes mmoX
, while the reductive dehalogenase genes vcrA
were recorded in 40 of 49 (82%) samples analysed (Table 2
, Figure 2
). All functional genes together (etnC
were detected in 38 of 49 (78%) samples analysed, indicating that both aerobic oxidation and reductive dechlorination of chloroethenes may take place simultaneously at the same place or in close microenvironments. This finding is consistent with the study of Liang et al. [20
], which detected functional genes from all three bacterial guilds (ethenotrophs, methanotrophs and reductive dechlorinators) in 99% of groundwater samples collected at six contaminated sites.
The qPCR results exhibited noticeable differences in individual biomarkers within one site or even one contaminant plume (e.g., aerobic biomarkers in contaminant plume #10_1 or reductive biomarkers in contaminant plume #1_1; see Table 2
Of the 49 analysed samples, 16 were collected during the on-going remedial biostimulation (samples were collected 1 to 4 months after the last whey application, see column 6 in Table 2
). Despite the fact that application of whey stimulates reductive dechlorination, all functional genes (etnC
coexisted in 14 of the 16 (88%) samples.
Of the anaerobic dechlorinators, Desulfitobacterium
spp. were most frequent, being present in 47 of 49 samples (96%), while D. mccartyi
spp. were identified in 46 of 49 analysed samples (94%). The occurrence of D. mccartyi
spp. corresponded well with Dehalobacter
spp. (see correlation analysis in Section 3.3
), though the latter were only identified in 74% of samples. The lower incidence of Dehalobacter
spp. in groundwater samples may reflect the fact that Dehalobacter
spp. degrade the parent chlorinated compounds PCE and TCE [3
], which were mostly degraded to less chlorinated metabolites at the sites tested. Frequent detection of Dehalogenimonas
spp. in our samples (94%) is consistent with the findings of Yang et al. [13
], who detected Dehalogenimonas
spp. in 81% of 1173 samples collected in the United States and Australia.
3.3. Correlation and Feature Selection Analysis
Correlation analysis (Figure 3
) revealed that TCE was only correlated with Desulfitobacterium
spp., probably as none of the enzymes tested related to the reductive or aerobic functional genes participating in biodegradation of TCE. However, cis
-1,2-DCE was positively correlated with both the reductive functional gene vcrA
and the ethenotroph functional genes etnC
, as well as the reductive dechlorinators Dehalogenimonas
spp. and Desulfitobacterium
spp. Similarly, VC concentration was positively correlated to the ethenotroph functional genes etnC
and the reductive dehalogenase genes vcrA
, as well as the reductive dechlorinators D. mccartyi
spp. The positive correlation of cis
-1,2-DCE and VC to abundance of Dehalogenimonas
spp., whose ability to degrade cis
-1,2-DCE and VC was only recently confirmed by Yang et al. [13
], indicates a potentially notable contribution of Dehalogenimonas
spp. in reductive dechlorination at the sites tested.
None of the chlorinated ethenes were correlated with the methanotroph functional genes mmoX and pmoA.
In sum, it implies that both anaerobic reductive dechlorination and ethenotroph-mediated aerobic biodegradation could participate in biodegradation of VC and cis-1,2-DCE.
With regard to the correlation of individual functional genes with hydrochemical and field parameters, the reductive dehalogenase genes vcrA
were negatively correlated with ORP and sulfate and positively correlated with dissolved iron, hydrogen sulfide and methane. As has been previously noted, anaerobic reductive dechlorination of chloroethenes can occur under both nitrate-reducing and iron-reducing conditions, though the most favourable conditions for anaerobic dechlorinators are sulfate reducing and methanogenic [43
]. This is also supported by a strong negative correlation of Cl no. with methane and positive correlation of Cl no. with sulfate. The methanotroph functional genes mmoX
were positively correlated with dissolved iron and methane, while pmoA
was negatively correlated with sulpfate. There was no correlation observed between the ethenotroph functional genes etnC
and any hydrochemical and/or field parameters, suggesting that hydrochemical conditions of aquifers are not limiting factors for proliferation of ethenotrophs.
The significance of individual cis
-1,2-DCE and VC biodegradation processes was assessed using feature selection utilising the random forest classification algorithm. Only the reductive dehalogenase gene vcrA
exhibited significant relevance for cis
-1,2-DCE (Figure 4
), whereas both reductive dehalogenase genes (vcrA
) and the ethenotroph functional gene etnE
were significantly relevant for VC (Figure 5
). In sum, feature selection provided further support for the hypothesis that both reductive dechlorinators and ethenotrophs participate in biodegradation of VC at the sites tested; however, it indicated that there was only intermediate significant involvement of ethenotrophs in biodegradation of cis
3.4. Hydrochemical Conditions for Aerobic Oxidation
As aerobic (both metabolic and co-metabolic) oxidation processes in remedial practice can overcome a frequent accumulation of metabolites cis
-1,2-DCE and VC generated during in-situ anaerobic reductive bioremediation, the hydrochemical conditions for these processes were assessed. First, the whole dataset was subjected to cluster analysis with the ethenotroph functional genes etnC
set as clustering variables. Six sample clusters were identified, with clusters #1 and #6 representing groundwater samples with the highest and lowest abundance of etnC
, respectively. The basic statistical parameters (maximum, mean, median, and minimum values) calculated for the hydrochemical parameters for each cluster are given in graphs in the supplementary material
(Supplementary Figure S1)
. The prevailing redox processes assessed for each groundwater sample (see Section 3.1
) are given in Supplementary Table S2
Cluster #1, which represented the highest potential for aerobic oxidation, was comprised of three groundwater samples collected from two sites. Redox conditions were mixed, covering a wide spectrum of processes from Fe(III) reduction (iron mean concentration 33.38 mg/L) to methanogenic (methane mean concentration 7.8 mg/L). The groundwater samples in cluster #1 differed from those in the other clusters, mainly in higher concentrations of VC (mean 10,314 µg/L) and of ethene (mean 4534 µg/L).
In comparison, cluster #6 comprised groundwater samples with lowest potential for aerobic oxidation and included 28 samples collected at eight sites. With regards to redox conditions, cluster #6 contained samples with Mn(VI) reduction and Fe(III) reduction predominating, along with samples of mixed redox categories, covering Mn(VI) reduction down to methanogenic conditions. Mean iron and methane concentrations for cluster #6 were 15.8 and 3.5 mg/L, respectively. Groundwater samples in this cluster contained significantly lower concentrations of VC (mean 276 µg/L) and ethene (mean 344 µg/L); however, the degree of chloroethene dechlorination (Cl no.) was similar to that of cluster #1 (mean Cl no. cluster #1 = 1.5, cluster #6 = 1.4). This indicates that aerobic oxidation of chloroethenes can take place in aquifers under a range of redox conditions, including apparently reducing environments. The most likely explanation is that the studied aquifers are not homogenous environments with one redox state only, but highly spatially and temporally heterogenous macro- and microenvironments with different redox conditions. Spatial heterogeneities may be a result of different lithology and permeability that influence migration patterns of organic substrates (electron donors), dissolved oxygen, chloroethenes and their co-occurring contaminants. This explanation is supported by the results of hotspot high-resolution characterisation performed on sites #2, #3, #5, #6, #7, #8 and 11 using a membrane interface probe (MIP) (Geoprobe Systems®, Salina, KS, USA). MIP was used for collection of semi-quantitative data on the presence of volatile organic compounds (VOCs) and soil electric conductivity measurements indicating aquifer lithology in the vertical soil profile. MIP profiles (data not shown) from majority of the sites show presence of soil layers with different electric conductivities and levels of contamination (i.e., distinct lithological and contamination heterogeneity).
Temporal changes may result from irregular seepage of oxic rainwater, groundwater level fluctuations, and/or from discontinuous application of organic substrates at the bioremediated sites. Methanotrophs are found mainly at aerobic/anaerobic interfaces in soil and aquatic environments that are crossed by methane [44
], thus inhabiting environments with low oxygen level. Similarly, ethenotrophs can survive in environments with very limited oxygen contents [17
]. These macro- and microheterogeneities and abilities of aerobic ethenotrophs to sustain low levels of oxygen are reasons for their occurrence in anaerobic aquifers containing anaerobic dechlorinators. Co-occurrence of anaerobic dechlorinators and VC assimilating bacteria was revealed also in groundwater samples from other sites [20
], in discrete aquifer soil samples [19
] as well as in surface riverbed sediment samples [47
] where the TOC content in soil was found to be the critical parameter determining the dominant degradation pathway [48
The results also suggest that high concentrations of ethene and VC (as electron donors in aerobic oxidation of chloroethenes) are correlated with the abundance of ethenotrophs. This is also supported by the results of correlation analysis, which indicated a positive correlation between VC and the ethenotroph functional genes etnC and etnE.