Trichloroethene (TCE) is commonly associated with various human activities, including its use as a metal detergent, dry cleaning agent, pesticide solvent, and chemical feedstock (DRC, 2012; Ottosen et al., 2020). TCE, along with its breakdown products, dichloroethene (DCE) and vinyl chloride (VC), are known to be teratogenic, carcinogenic, and mutagenic, posing serious threats to human health (Ordaz et al., 2017). Microbial anaerobic reductive dechlorination is an effective method for eliminating TCE from water systems. This process sequentially converts TCE to DCE, then to VC, and finally to non-toxic ethene. Many dechlorinating bacteria, including species from Proteobacteria (such as Geobacter, Desulfomonile, and Desulfuromonas) and Firmicutes (such as Dehalobacter and Desulfitobacterium), can partially catalyze the reduction of TCE to DCE, while Dehalococcoides mccartyi and Dehalogenimonas, which belong to the phylum Chloroflexi, can completely reduce TCE to ethene (Löffler et al., 2013; Yang et al., 2017). However, the effective use of Dehalococcoides to bioremediate the sites contaminated with chlorinated ethenes is limited to several factors, such as pH and the presence of co-existing electron acceptors.

The pH value significantly affects the TCE reductive dechlorination process. The pH range of Dehalococcoides is between 6 and 8, with the highest activity occurring between 6.9 and 7.5 (Löffler et al., 2013; Asai et al., 2022).

When the pH of the medium was 6.0, the dechlorination activity of Dehalococcoides decreased, leading to the accumulation of more toxic intermediate metabolites (DCE and VC). At a pH of 5.5, the dechlorination activity was lost (Yang et al., 2017; Ortiz-Medina et al., 2023). Reductive dechlorination and fermentation of organic substrates are acid-producing processes that lead to a decline in aquifer pH values (Brovelli et al., 2012; Robinson et al., 2009). Therefore, buffering techniques, including alkali addition and the use of non-acidified electron donors, are implemented to adjust and control aquifer pH to ensure the successful in situ bioremediation of contaminated sites (Borden et al., 2016; Lee et al., 2024). Sodium bicarbonate, or sodium carbonate, is often used as a buffer in the medium, but high bicarbonate concentrations can promote the growth of autotrophic methanogens and acetogens and consume large numbers of protons, resulting in an increase in the pH of the system. For example, when 10mM of HCO₃^– was added to a medium with an initial pH of 7.5, the final pH rose to 8.7 (Borden et al., 2016; Delgado et al., 2012; Steffan and Environmental Security Technology Certification Program, 2010). In addition, insoluble silicate minerals and colloidal Mg(OH)₂ were tested as long-term sources of alkalinity to control pH, while silicate minerals showed an inhibitory effect on the reductive dechlorination of DCE and VC. Conversely, colloidal Mg(OH)₂ was successfully employed to control groundwater pH for over 5 years (Hiortdahl and Borden, 2014; Lacroix et al., 2014; Ortiz-Medina et al., 2023). Furthermore, sodium formate was used as an electron donor to minimize acidification, and a blend of emulsified substrate and modified rice husk ash was used as a green and long-lasting substrate (GLS) for slow carbon release and pH control (Lee et al., 2024; Philips et al., 2013; Tomita et al., 2022). However, the use of formate and GLS as pH buffers may be incapable of neutralizing H⁺ produced by high dechlorination rates because of the slow-release process. While exogenous substrates are added, natural buffering processes such as nitrate reduction also affect aqueous pH. However, their influence on reductive dechlorination remains unreported.

Nitrate (NO₃⁻), a frequent co-contaminant with TCE, is mainly used in rocket propellants, explosives, and agricultural fertilizers (Carrey et al., 2021). In China, the highest nitrate concentration was up to 265 mg/L, and the detection rate of TCE was approximately 12% in some urban areas because of the unreasonable treatment of petroleum and textile pollution and the abuse of pesticides and fertilizers (Liu et al., 2017). On a global scale, the detected concentration of TCE ranged from 110 to 310 μg/L, and the maximum contaminant level of nitrate in the majority of wells was between 20 mg/L and 45 mg/L in the Fresno-Clovis metropolitan area (Kloos, 1997). In some industrial areas in Canada, such as Barrie-DC, Halifax Regional Municipality, and Greenwood, the detected concentrations of TCE and nitrate can reach up to 41 mg/L and 35 mg/L, respectively (Roy and Bickerton, 2012). NO₃⁻ can be reduced to N₂ through denitrification by denitrifying bacteria with organic matters (such as methanol, acetic acid, ethanol, etc.) or inorganic matters (such as hydrogen (H₂), elemental sulfur, zero-valent iron, etc.) as electron donors (Rittmann and McCarty, 2001; Zhao et al., 2011). In the entire denitrification process, 5 moles of electrons are required for 1 mole of NO₃⁻ reduction. There are many types of denitrifying bacteria, including more than 50 genera and 130 species, which are distributed among α, β, γ-Proteobacteria, Deferribacteres, Firmicutes, and Actinobacteria (Stein and Klotz, 2016; Zhao et al., 2013; Zhong et al., 2017).

The denitrification process is a self-producing alkali process, and 1 g of nitrate or nitrite consumption produces 3.57 g of basicity (calculated as CaCO₃). In the high-load denitrification process, the pH value of the reaction solution can easily exceed 9.0 and inhibit the activities of denitrifying functional bacteria (Li et al., 2016; Shi et al., 2023). Many studies have found that nitrate inhibits TCE dichlorination, and the higher the nitrate concentration, the stronger the inhibitory effect (Cao et al., 2017; Xu et al., 2018; Zhang et al., 2022). Yin et al. (2018) verified that the denitrification byproduct N₂O reacted with the super-reduced Co(I)-corrinoids of reductive dehalogenases to strongly inhibit the dechlorination activity. However, the pH adjustment by the interaction between nitrate self-alkalization and TCE self-acidification, along with their effects on nitrate reduction and reductive dichlorination, has not been studied.

This study established a series of experiments to evaluate the interaction between TCE and nitrate with or without external buffering substrates in mixed anaerobic dechlorinating consortia. TCE reductive dechlorination efficiencies, nitrate reduction, pH variation, and bacterial community structures were thoroughly investigated. The objectives of the present study were: (i) to explore the simultaneous reduction of nitrate and TCE under different pH conditions; (ii) to determine the effect of nitrate self-alkalization on the reductive dechlorination efficiency; and (iii) to reveal the changes in microbial community composition.

To evaluate the interaction between TCE and nitrate, we conducted a set of experiments in serum bottles. The serum bottles contained microbial consortia that were capable of reducing TCE and/or nitrate. The bottles were amended with lactate as an electron donor and TCE and nitrate at different initial concentrations. To achieve the three objectives, we measured the concentrations of TCE, nitrate, and pH values during the experiments. We also studied the microbial community structure to further understand the interaction between TCE and nitrate from a microbial perspective. The experiment details are described below.

We conducted a series of experiments to investigate the interaction between TCE and nitrate bioreduction under varying pH conditions. This information is essential for the bioremediation of chlorinated ethenes and nitrate, which are generally present as co-contaminants in contaminated sites. We found that nitrate reduction affected reductive dechlorination relatively differently under pH conditions without external buffering agents.

The transformation of TCE to cis-DCE was inhibited by nitrate reduction, which is similar to previous studies. For example, Fallahpour et al. (2018) demonstrated that TCE remediation efficiency dropped by approximately 10% in the presence of 40 mg/L (0.47 mM) nitrate compared to 2 mg/L (0.02 mM). Similarly, Zhang et al. (2022) reported that TCE removal efficiency decreased from 95.28 ± 0.4% to 85.98 ± 0.2% when the nitrate concentration increased from 25 mg/L (0.29 mM) to 100 mg/L (1.18 mM).

This inhibition of TCE degradation to cis-DCE may be attributed to the toxic effects of nitrate or denitrification byproducts, nitrite, and N₂O, on dechlorinating populations. Both nitrite and N₂O are known inhibitors of reductive dehalogenases and have been confirmed to strongly inhibit dechlorination activity (Neumann et al., 1995; Suyama et al., 2001; Yin et al., 2018).

Cao et al. (2017) also indicated that nitrate strongly inhibits the reductive dechlorination of 2,4,6-trichlorophenol, while Boopathy and Peters (2001) observed that under nitrate-reducing conditions, reductive dechlorination ceased at cis-DCE and trans-DCE. However, the conversion of cis-DCE to VC was not significantly affected in this study. The inhibitory effect of nitrate on dechlorination diminished once it was fully reduced (Cao et al., 2017; Zhang et al., 2022).

Notably, contrary to previous findings, we found that the conversion of VC to ethene was enhanced by nitrate reduction. This enhancement may be linked to pH changes associated with nitrate reduction. In mixed dechlorinating consortia, processes such as reductive dechlorination and lactate fermentation produce H⁺, which acidifies the solution. In contrast, nitrate reduction is a self-alkalizing process that alkalizes the solution.

In our experiments, TCE was added as the sole electron acceptor without buffering substrates. The pH in cultures without nitrate gradually decreased due to fermentation and reductive dechlorination, leading to a decline in the TCE reduction rate. This decrease is critical as Dehalococcoides, the genus responsible for converting VC to ethene, is highly sensitive to pH, with an optimal growth range of 6.9 to 7.5 (Lee et al., 2024; Yan et al., 2021; Yang et al., 2017). In cultures with added nitrate, the pH increased as nitrate was reduced. This effect was particularly pronounced at a nitrate concentration of 3 mM, where the pH increased to 7.75 by the end of the experiment. These findings indicate that nitrate can function as a buffer to stabilize pH and enhance dechlorination rates. However, a nitrate concentration of 0.3 mM proved to be more effective than 3 mM for promoting the dechlorination process. This highlights the potential for leveraging nitrate self-alkalization in optimizing bioremediation strategies.

In situ, contaminated sites often exhibit acidic conditions due to natural factors such as high concentrations of dissolved carbon dioxide, desorption of H⁺ from clay minerals, or human-influenced causes such as acid rain (Ortiz-Medina et al., 2023). In regions with higher rainfall, pH values below 6 are commonly observed (Zhou et al., 2015).

To further explore the potential of nitrate self-alkalization to adjust pH under suboptimal conditions, simultaneous reduction experiments were conducted using NaHCO₃ and tris-ethanesulfonic acid as buffers to maintain the initial pH below 6.5. We found that the reduction of 0.3 mM nitrate effectively mitigated the pH decrease and enhanced the reduction rate of VC to ethene. However, at a nitrate concentration of 3 mM, despite the pH reaching the optimal growth range, the dechlorination process was significantly inhibited, leaving 74.73% DCE and 25.27% VC as the final products.

These findings confirmed that nitrate reduction can increase pH even in the presence of buffering substrates. Under suboptimal pH conditions, a moderate nitrate concentration (e.g., 0.3 mM) slightly promoted the reductive dechlorination process. However, higher nitrate concentrations (e.g., 3 mM) severely inhibited dechlorination, emphasizing the importance of optimizing nitrate dosage for effective bioremediation.

As expected in the dechlorinating consortia, we observed nitrate reduction and methanogenesis. Methanogenesis was strongly inhibited in this study, which indicated that 0.3 mM TCE is toxic to methanogens. Various studies have shown that chlorinated ethenes are toxic to microorganisms. TCE (10 mg/kg) inhibited the process of nitrate transformation into nitrite by affecting the activity of nitrate reductase in Mollisol (Li et al., 2020). However, we found the tested TCE concentration had no inhibitory effect on the nitrate reduction, and the nitrate removal rate was faster in the presence of TCE than in the culture with nitrate as the sole electron acceptor. This may be due to the inoculated culture used in this study, which had never been exposed to nitrate before and needed an acclimation time to activate the process of denitrification.

Further electron distribution study showed that the electrons flowing to TCE reduction declined when the nitrate concentration increased. On the one hand, denitrifying bacteria are prevalent in the environment and have diverse metabolic pathways (Sun et al., 2024; Zhang et al., 2022). On the other hand, nitrate is a more thermodynamically favorable electron acceptor than TCE or its daughter products (Table 2), so the reductive dechlorination process can be easily outcompeted.

Based on the sequencing results, we found that Dehalococcoides belonging to Chloroflexi and putative acetogenic bacteria belonging to Clostridium_sensu_stricto_7, Lentimicrobium, and Spirochaetaceae dominated the consortia added with TCE as an electron acceptor. Many studies have shown the co-existence of Dehalococcoides and these fermenters, which stimulated Dehalococcoides growth by providing acetate and corrinoid cofactors (Wen et al., 2020; Yan et al., 2021). However, we found that the abundance of Dehalococcoides and Clostridium declined in cultures with higher concentrations of nitrate, revealing the inhibitory effect of nitrate on Dehalococcoides and Clostridium. While Lentimicrobium maintained a relatively high abundance in the presence of nitrate in this study. Lentimicrobium has been shown to grow in the pH range of 6.6–9.0 and has been shown to play a significant role in the reduction of high concentrations of nitrate (Li J. et al., 2022; Li Y. et al., 2022; Sun et al., 2016; Wang et al., 2020; Zhang et al., 2024). In comparison, we found that Petrimonas and Trichlorobacter dominated the cultures exposed to TCE and nitrate. Chen et al. (2023) showed that hydrogen production and DCE removal were positively correlated with Petrimonas. Grabowski et al. (2005) reported that Petrimonas can use nitrate as an electron acceptor.

Similarly, we found that the abundance of Petrimonas increased, especially when the nitrate concentration was up to 3 mM (255 mg/L). This implied that Petrimonas was potentially involved in the process of dechlorination and denitrification in the culture exposed to high concentrations of nitrate. Trichlorobacter was positively correlated with TCE dechlorination efficiency (Chen et al., 2023; De Wever et al., 2000). We found that the abundance of Trichlorobacter was negligible in the culture with TCE as the sole electron acceptor, but was dominant in the culture in the presence of nitrate. This indicates that Trichlorobacter is possibly responsible for nitrate reduction and outcompetes Petrimonas when nitrate concentrations are low. Much more research is needed to understand the mechanism of nitrate action at TCE-contaminated sites further.

We evaluated the interaction between nitrate and TCE through a series of experiments conducted in serum bottles. The results indicated that TCE concentrations up to 0.3 mM promoted nitrate reduction, while stepwise reductive dechlorination was differentially affected by nitrate under varying pH conditions. Although nitrate inhibited the reduction of TCE to DCE, an appropriate nitrate dose enhanced the conversion of VC to ethene. This outcome was mainly due to the thermodynamical preference of nitrate reduction compared to TCE reduction, while the enhancement of VC conversion was linked to pH stabilization by the self-alkalization process.

As the nitrate concentration increased from 0 to 3 mM, the relative abundance of fermentative Clostridium abundance sharply decreased from 31.11 to 1.51%, whereas putatively denitrifying genera Petrimonas and Trichlorobacter showed significant increases. Furthermore, when the nitrate concentration was increased to 0.3 mM, the relative abundance of Dehalococcoides slightly increased from 23.91 to 24.26%. However, at a nitrate concentration of 3 mM, Dehalococcoides abundance decreased to 18.65%, indicating a tolerance to high nitrate concentrations under specific conditions.