1 Introduction

Trichloroethene (TCE) and arsenic (As) are two of the most prevalent groundwater pollutants, and they often coexist persistently in soil saturated zone (Podgorski and Berg 2020; Zhang et al. 2022). The U.S. Geological Survey reported that TCE existed in 5% groundwater (Moran et al. 2007), and as one of the most frequent co-contaminants with TCE, As exists at approximate 63% of current TCE-contaminated National Priorities List sites (Gushgari-Doyle and Alvarez-Cohen 2020). An example in Shanghai, China is the contaminated groundwater in Taopu industrial park containing 40–50 mg L−1 TCE and 1–2 mg L−1 As, which come from a nearby pharmaceutical factory. Exposure to TCE is associated with many health issues such as altered immune function, cancer, or disrupted systemic metabolism (Tachachartvanich et al. 2018), while people drinking As-containing water suffer a risk of visceral cancers (Zhang et al. 2019). In the groundwater environment with low redox potential, TCE and As undergo a gradual reduction transformation (Islam et al. 2021; Kumar et al. 2020). Part of arsenate [As(V)] is reduced to arsenite [As(III)] with higher toxicity (Gushgari-Doyle and Alvarez-Cohen 2020). Intermediate products such as dichloroethene (DCE) and vinyl chloride (VC) are produced slowly through natural biodegradation, which are less biodegradable, more toxic, and more mobile than TCE (Scheutz et al. 2011). It is difficult to achieve complete dechlorination by natural processes, and thus these intermediate products become new groundwater contaminants threatening environment and public health (Deng et al. 2021).

Despite  the huge application potential, cost-effective and environmentally friendly bioremediation technologies for the simultaneous removal of TCE and As are yet to be developed and deployed at scale, as such approaches are extremely sensitive to environmental conditions and pollutant types/concentrations (Gaur et al. 2018; Gaza et al. 2019; Sheu et al. 2018). For the combined contamination of TCE and As, it’s difficult to achieve simultaneous removal via a single microbial strain because of the fundamentally different physicochemical properties of the contaminants. While numerous microorganisms have demonstrated the ability to reduce TCE and its intermediate products, only Dehalococcoides mccartyi and Dehalogenimonas are capable of converting TCE to the benign product ethene via reductive dechlorination (Polasko et al. 2019; Puentes Jácome et al. 2019; Yan et al. 2021; Yang et al. 2017). The metabolic activity of microorganisms is likely inhibited by the toxicity of pollutants and sometimes even by their intermediate products and byproducts. Haest et al. found that KB-1 culture containing Dehalococcoides spp. showed the highest biodegradation activity at 0.3 mM TCE, while no activity was found as TCE concentrations increased to 4–8 mM (Haest et al. 2010). Also, Sara et al. demonstrated that 9.1 μM As(III) could cause a 50% decrease in Dehalococcoides mccartyi cell growth, and thus TCE dechlorination would be inhibited (Gushgari-Doyle and Alvarez-Cohen 2020). Some bacteria are known to possess mechanisms for either oxidizing As(III) or reducing As(V), including transformations linked to energy generation. As(V) respiratory reductase (arr) genes encode periplasmic As(V) reductase that functions in anaerobic respiration using As(V) as a terminal electron acceptor, while aox genes encode periplasmic As(III) oxidase that transforms As(III) to As(V) (Liao et al. 2011). Despite the ability of these microorganisms to tolerate As, either by self-detoxifying or metabolizing it for energy generation (Jebelli et al. 2018; Yan et al. 2019), their effectiveness is still largely dependent on pollutant concentrations (Viacava et al. 2020).

Biochar (BC), as a carbonaceous material derived from pyrolysis of biomass at high temperatures under O2-starved conditions, owing to its relatively high surface area, multilevel porous structure (Huang et al. 2020), and electron donor capacity (Zhang et al. 2018), has been reported to enhance the microorganisms’ adaptability to adverse environments in many bioremediation systems (Zhao et al. 2020). It is illustrated that BC facilitated chemical interactions between organic pollutants and organic degrading microorganisms colonizing biochar surfaces, enabling effective biodegradation (Patel et al. 2022). In addition, BC can protect the microorganisms from fatal acid inhibition due to its alkaline property (Lü et al. 2016). Biochar addition was also reported to enhance the structural integrity and electron-transfer capability of anammox sludge (Xu et al. 2022). Our earlier research demonstrated that BC shortened the time for 100% removal of TCE from 360 to 168 h and promoted selective colonization of reductive dechlorinating microorganisms on BC surface (Liu et al. 2021). However, pristine (unmodified) BC cannot adsorb As (AsO3−, AsO43−), and adversely, it results in a higher mobility of As due to the negative charges on BC (Amen et al. 2020). Many researches used Fe-modified biochar to immobilize As (Wen et al. 2021; Yang et al. 2022).

Until now there is neither study investigating BC’s effect on TCE removal in the presence of As, nor study trying to simultaneously remove these two contaminants via BC-mediated biological process. This study aims to investigate if biochar could strengthen biodechlorination in the presence of As, and further achieve simultaneous removal of TCE and As from groundwater. The intermediate biodechlorination products and As species were measured both in aqueous solution and on solid biochar. Desorption experiment of the organic contaminants from BC was conducted to elucidate the BC’s roles in regulating biodechlorination. Metagenomic and metabolomic analyses were performed to reveal the responses of microbial flora, functional genes and metabolites to BC addition. This work for the first time gives a systematic insight into interactions of biochar-microorganisms-coexisting contaminants, and is innovative to remove As efficiently through the filtration of As-adsorbed microbes by biochar. This study would promote the development of fast and complete bioremediation processes for co-contaminated groundwater.

2 Materials and methods

2.1 Biochar selection

A waste biomass, peanut shell, was selected as precursor to produce biochar due to its porous and fibrous structure. The detailed preparation process and the physicochemical properties of the biochars including elemental content, specific surface area, electron donating/accepting capacity, etc. can be found in our earlier publication (Liu et al. 2021). Based on the previous studies, BC’s promoting effect on biodechlorination did not vary greatly with biochar production temperature (300 °C, 500 °C, and 700 °C) during the early stage of the process, while at the later stage, BC produced at 700 °C (BC700) was more beneficial for the microbial process due to its largest specific surface (22.42 m2 g−1) and highly carbonized structure (Liao et al. 2011; Liu et al. 2021). Therefore, BC700 was selected for experiments in this study.

2.2 Simulated co-contaminated groundwater

Two different levels of contamination in the simulated groundwater, i.e., 10 mg L−1 TCE with 1 mg L−1 As(V) and 30 mg L−1 TCE with 4 mg L−1 As(V), respectively, were selected to study the remediating ability of microbes under different toxic stresses both of which matched matching with the real pollution conditions and our microflora’s co-removal ability (Liu et al. 2021). These low and high initial concentrations were referred as “LC” and “HC” samples. We prepared the TCE and As co-contaminated groundwater by firstly adding TCE into methanol to obtain a stock solution of 1000 mg L−1 (Wu et al. 2018), and then taking a certain amount of this stock TCE solution to mix with sodium hydrogen arsenate heptahydrate and added into deionized water to obtain the diluted TCE and As(V) solution with the set concentrations. Here methanol acted as a co-solvent for promoting the dissolution of TCE, and its content was 0.002% and 0.006% (v/v) for the LC and HC system, respectively. Dissolved oxygen in the solution was removed by nitrogen sweeping from the closed system to ensure anoxic conditions. To ensure adequate carbon source and electron donor for the subsequent biodechlorination, 10% yeast extract and 60% sodium lactate were added into the groundwater with the rate of 0.1% and 0.072%, respectively (Wang and Wu 2017).

2.3 Co-removal experiments

Batch experiments were performed in sealed glass serum bottles, and no headspace remained to ensure that all the volatile substances produced during the removal process were retained in the solution. The microbial suspension with OD600 being 0.5 indicating 107–108 cell mL−1, was inoculated into the co-contaminated groundwater by 4% (w/v). The dechlorinating microflora was obtained from a long-term running TCE anaerobic degradation system originally inoculated with activated sludge. The detailed adaptation procedure is shown in Additional file 1: Section S1. BC700 was amended into the system with the addition rate of 0.2% (w/v). We also setup several controls without microbial inoculation or without BC, as well as the blank control without any additives. In the sterilized control, we ruled out the bacteria using autoclaving combined with the filtration of microorganisms by 0.22 μm filter (Zhang et al. 2021). All serum bottles were cultured in incubator at 25 ± 0.5 °C without light for 18 days, and each trial was performed in triplicate. The experimental set-up and analysis items are listed in Additional file 1: Table S1. The mixture was sampled  on days 0, 1, 3, 6, 9, 12, 15, and 18 to measure the contaminants in the aqueous solution after solid–liquid separation.

2.4 Adsorption experiment and desorption experiment

To distinguish the role of biochar adsorption for these two contaminants and the microbial effects during their removal, we solely conducted separate TCE and As(V) adsorption experiments with biochar without inoculating microbes. The treatment conditions were consistent with those in the above co-removal experiment. The data were fitted to pseudo-first-order and pseudo-second-order equations, as well as Elovich and intraparticle diffusion model.

In order to clarify the TCE distribution in the aqueous solution (TCEaq) and on BC (TCEs) during the process, desorption of the TCE from BC was performed based on an extraction method referring to NIOSH protocol 1022 (Siggins et al. 2021), and the extraction efficiency for TCE was 84.11%. Briefly, 0.1 g BC was added into 5 mL carbon disulfide containing 2% (v/v) 2-phenoxyethanol and then mixed in a shaker for 30 min. The solution was filtered through a 0.45-μm filter membrane. The desorption of intermediate products including cis-1,2-DCE, 1,1-DCE and VC was also performed with this procedure, and their extraction efficiencies were 70–85%.

2.5 Measurement of trichloroethene and intermediate products

The concentrations of organic contaminants including TCE, cis-1,2-DCE, 1,1-DCE and VC were determined using the method of HJ 620-2011 in Ministry of Ecology and Environment of the People’s Republic of China. A headspace gas chromatography with a capillary column (60 m in length× 0.25 mm in diameter × 1.4 μm film) and an electron capture detector was used (7890B, Agilent Technologies Inc., USA). More detailed information can be seen in Section S2. To further verify the dechlorinating process, the change of Cl− in solution was measured by ion chromatograph (MIC, METROHM, Switzerland).

2.6 Determination of arsenic concentration and speciation

The As transformation and distribution in aqueous/solid phase were determined. The total As concentration in aqueous phase (Asaq) was determined using an inductively coupled plasma optical emission spectrometer (iCAP PRO, Thermo Scientific, USA), while the soluble inorganic As [As(III)aq and As(V)aq] was measured using high-performance liquid chromatography coupled with atomic fluorescence speciation analyzer (HPLC-AFS, Jitian, China). The organic As in the aqueous phase including monomethylarsonic acid (MMAA) and dimethyl-arsinic acid (DMAA) was also detected qualitatively using the HPLC method. The extraction of inorganic As in solid phase [As(III)s and As(V)s] was conducted based on GB 5009.11-2014 (National Health and Family Planning Commission of the People’s Republic of China, 2015). Detailed procedures are presented in Section S3. The extraction efficiencies for total As, As(III)s and As(V)s were 86.25%, 72.14% and 76.16%, respectively.

To identify the further methylation of As(III) (Soares Guimarães et al. 2019), a novel technique based on silver nitrate impregnated silica gel filled tubes was used to determine whether arsines [including arsine (AsH3), monomethylarsine (MeAsH2), dimethylarsine (Me2AsH) and trimethylarsine (TMAs)] were produced (Mestrot et al. 2009). The designed equipment consisting of arsines-trap tubes, peristaltic pumps and the co-removal system is shown in Additional file 1: Figure S1. The customized polymethyl methacrylate tube was loaded with 10 g 10% AgNO3 impregnated silica gel. The whole system was airtight, and the peristaltic pump pushed the gas generated by the microbes through the impregnated silica gel traps for the whole duration of the experiment. Traps were then eluted by hot boiling nitric acid for measuring the total As.

2.7 Metagenomics sequencing and metabolomics analysis

To study the temporal distribution of microbial colony, we collected the microbial samples from the different stages on day 6 and day 15, respectively, labeled as ES (early stage) and LS (late stage). In the systems with BC addition, microorganism colonies attached to the biochar surface and suspended in the aqueous solution were collected separately and named “BS” and “AS” as “BS” and “AS”, respectively. Specifically, 100-μm cell strainer and 50-mL centrifuge tube were assembled to achieve the separation of microorganisms existing in different phases because the size of prepared BC was bigger than 100 μm while microorganisms were much smaller. The entire filtration process was conducted in the larminar flow. The microbial samples obtained from the systems without biochar were marked as “NB”. Therefore, we obtained 12 microbial samples for metagenomics sequencing. They were named as LC-NB-ES, LC-NB-LS, HC-NB-ES, etc. For instance, LC-NB-ES represented the samples collected from the “No BC” system with low contamination concentrations in the early stage. For clarity, all the abbreviations and full names are listed in Additional file 1: Table S2, and we also labeled them clearly in the figures. To perform the metabolomics analysis, 8 groups of supernatant samples were obtained by high speed refrigerating centrifugation (24,000 rpm, 4 °C) of the mixture. Trials were performed in triplicate and all samples were sent to Shanghai Majorbio Bio-pharm Technology Co. Ltd. for sequencing analysis. The detailed molecular biological analysis methods can be found in Additional file 1: Section S4.

3 Results and discussion

3.1 Simultaneous dissipation of trichloroethene and arsenic

The removal kinetics of TCE and total As from the synthetic groundwater are shown in Fig. 1. The TCE removal rate was calculated based on the measured TCE reduction rather than on the observed dechlorination products. For the blank control without microbes and BC, the two contaminants almost did not change in 18 days treatment. In contrast, the trial of microbial inoculation with no BC addition showed a total TCE removal of 85.72% and 75.13%, respectively, for the low and high initial TCE concentrations on the 18th day. The addition of BC resulted in a remarkable acceleration in the TCE removal and increased the total removal rate, i.e., a complete removal was achieved on the 9th day and the 12th day for LC and HC systems, respectively (Fig. 1a). Meanwhile, more obvious difference was observed in As removal process between No BC and BC-amended systems. That is, 50–70% of As was removed from the solution within approximate 6 days with the amendment of BC, while the No BC systems showed a similar pattern to the blank control, i.e., nearly no As was removed throughout the process nearly no As was removed throughout (Fig. 1b).

Fig. 1
figure 1

Removal dynamics of TCE (a) and total As (b) from the contaminated groundwater by microbes with the assistance of biochar. LC: Low concentration of contaminants, 10 mg L−1 TCE with 1 mg L−1 As(V); HC: High concentration of contaminants, 30 mg L−1 TCE with 4 mg L−1 As(V)

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In the system without BC, there was a clear lag of 6 days before the start-up of the fast TCE decrease, while in BC-amended system, TCE concentration began to decrease from the start and was removed completely without any lag (Fig. 1a). To our surprise, TCE degradation with BC assistance was even faster in the high TCE concentration than in the low TCE system with no BC, indicating that the promotion of biodegradation by BC was particularly effective for high initial concentrations of contaminants (Fig. 1a). As for the arsenic removal, a nearly vertical decrease curve was observed on day 1 in the trials with microorganisms and the addition of BC, and then gradual decrease of As concentration occurred from the 1th day to the 12th day, followed by subsequent slight fluctuations (Fig. 1b). Therefore, solid–liquid separation should be conducted on day 12 to simultaneously achieve the highest As removal and the complete TCE removal.

The above results showed that BC amendment not only enhanced the overall TCE removal efficiency, but also significantly accelerated the degradation rate through shortening the microbial lag phase and strengthening the activity of microorganisms. It was notable that during the whole process, the removal rates of the LC systems were higher than those of the HC systems no matter whether BC was present or not, which indicated that high concentrations of contaminants had an inhibiting effect on microbes. Our previous studies demonstrated the lethal toxicity of higher concentrations of organic contaminants to bacteria (Liu et al. 2021; Zhao et al. 2020). In this study, the used inoculated bacteria used were taken from a long-term running TCE anaerobic degradation system, and thus the microorganisms had relatively high resistance to contaminant toxicity. Moreover, the results showed that BC presented a significant promoting effect on As removal (Fig. 1b). Therefore, biochar acted as a carrier to separate the microorganisms from As adsorbed from the groundwater (Additional file 1: Figure S2), and this method could be easily  implemented in actual remediation process. Microorganisms played major roles in removing As directly, while the BC’s function lied  in separating the microorganisms from As in them from the groundwater.

3.2 Accelerated complete biodechlorination by biochar with coexistence of arsenic

In the system with both biochar and microorganisms, adsorption and biodegradation occurred simultaneously. To elucidate how BC assisted the microbes to achieve the fast and complete biodechlorination, we conducted sole adsorption experiment of TCE to BC without inoculation (Additional file 1: Table S3), which could distinguish the BC’s roles in removing TCE via adsorption or accelerated biodegradation. Moreover, the residual TCE on BC was desorbed to illustrate the degradation dynamics of TCE in aqueous phase and on BC, as well as their spatial distributions. TCEaq was removed completely on day 9 in LC system with BC, while 3.61 mg L−1 TCEs still existed and declined to 0.16 mg L−1 on day 18 (Fig. 2a). That meant the almost complete biodegradation of TCE required 18 days, while faster removal from the aqueous solution could be achieved within 9 days by picking out the TCE-adsorbed BC. However, a small portion of TCEs (4.08 mg L−1) still remained in the HC system (Fig. 2b).

Fig. 2
figure 2

Concentration change of the dechlorination-related products in the co-removal systems. a and b refer to the gradual decrease of TCE distributed in aqueous (TCEaq) and biochar solid phase (TCEs), respectively, in the low concentration (LC) and high concentration (HC) systems; Change curves of the adsorbed vinyl chloride (VC) on the biochar solid phase and the produced Cl− are shown in c and d, respectively

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According to data of the sole adsorption experiment by BC700, TCE achieved the highest adsorption capacity (36.11% in LC system; 29.87% in HC system) within 48 h (Additional file 1: Figure S3a). This was regarded to be responsible for alleviating the exposure toxicity of TCE to microbes, assisting them to grow against the adverse environment, even survive in high organic contaminant concentrations with fatal toxicity (Liu et al. 2021). However, it was notable that even with adsorption of 29.87% of TCE by BC700 in the HC system, the residual TCE concentration was still 21.04 mg L−1 (initial TCE concentration was 30 mg L−1, i.e., much higher than the initial TCE concentration in the LC system (10 mg L−1). Despite this, there was no lag observed for the onset of microbial degradation (Fig. 1a). This strongly suggested that BC’s role in promoting biodegradation was not limited to the TCE adsorption, alleviating toxicity to microbes.

We tried to detect the intermediate products such as 1, 1-DCE, cis-1, 2-DCE and VC, and we obtained their standard curves (R2 > 0.999) (Additional file 1: Figure S4). However, we did not detect these intermediate products in all solutions, and only observed the release of gaseous final product. Only the BC-adsorbed VC was detected throughout the removal process, and the fluctuations of VC concentration on BC interestingly reflected the dynamic adsorption and degradation processes. There existed a sharp VC increase between days 1 and 3 corresponding to the sharp decrease of TCE, and then an obvious decrease of VC followed during days 3–6 with its subsequent gradual removal (Fig. 2c). The initial accumulation of VC was due to the slight biodegradation at the initial stage, while the biodegradation became vigorous over time. This was in agreement with an observation that as the number of chlorine substituents decreases, dechlorination rate usually slows down with accumulation of less chlorinated DCE and VC (Hermon et al. 2019). Finally, almost no intermediate products remained in the dechlorinating process.

The end products, ethylene and chloride ion (Cl−) were detected. Ethylene had been detected in our former microflora domestication experiments with gaseous head space. The results presented in Fig. 2d clearly showed that Cl− concentration increased with time except for the blank control, and along the whole process it was much higher in the systems with BC incorporated than the trials without BC addition. Furthermore, this enhanced-production of Cl− in presence of BC was more significant in the HC system than in the LC system.

From these results it can be concluded that complete biodechlorination could be achieved in this mixed microorganism system without accumulation of intermediate products both with and without BC addition. However, BC can significantly accelerate the TCE degradation via adsorbing the intermediate products such as VC, which is more toxic than the starting contaminant, enabling more TCE to be converted into end products in a shorter time. This promoting effect was particularly remarkable in the high concentration pollutant system (Fig. 2d).

3.3 Preferential colonization of dechlorination bacteria on biochar

To study the influences of BC on the dynamic changes of microorganisms responsible for TCE dechlorination, relative abundances of OTUs at different taxonomic levels were calculated (Koner et al. 2022).  At the genus level, four categories of TCE-degradation related microorganisms were dominant in the systems (Fig. 3). The crucial reductive dechlorinating microorganisms (RDM: Clostridium and Dehalococcoides) were found to exist mainly on BC surface. On the contrary, the co-metabolic dechlorinating microorganisms (CDM: Comamonas and Pseudomonas) preferred to suspend in solution, and thus it can be inferred that BC addition caused spatial differences in this system regarding to microbial community structure, substrate-metabolism, etc. These differences enabled the RDM and CDM to proliferate selectively on biochar surface and in liquid solution, respectively, to realize the coordination of space utilization and “niche differentiation”. This phenomenon indicated that there was a large difference between microbial communities in the liquid and on biochar surfaces for different dechlorinating activities. The auxiliary dechlorinating microorganisms (ADM: Exiguobacterium, Trichococcus and Sedimentibacter) showed a similar distribution pattern as RDM, likely assisting RDM. For example, Trichococcus produces electron donor (lactate and acetate) for TCE dechlorination through glucose fermentation (Zhang et al. 2015). In addition, in HC system, the relative abundance of Clostridium was clearly higher on BC surface of the BC amended system than in aqueous solution of No BC system, while the opposite was true for the hydrogen-competing microorganisms (HCM: Desulfovibrio) counting against RDM’s dechlorination activity. It was attributed to the massive reproduction of RDM on BC surface maintaining the hydrogen concentration at a low level which was adverse for HCM’s growth (Kushkevych et al. 2021). It is important to note that, although being adverse to dechlorination, the relative abundance of Desulfovibrio was the highest among all the microbes assessed, owing to its function related to the most crucial As(V) reduction process, and thus BC’s suppression on this HCM was important (Angai et al. 2022; Xiu et al. 2021). Source data are provided in Additional file 1: Figure S5, S6 and S7.

Fig. 3
figure 3

Heatmap of the top 20 abundant genera (excluding the genus undefined at the genus level). Darker and lighter blue indicate the relative abundance of each OTU in different samples depending on Z-score (Z = (x − μ)/σ, μ is sample average, σ is standard deviation of samples). LC: Low concentration; HC: High concentration; NB: Microbial samples in “No BC” systems; AS: Microorganism suspended in the aqueous solution; BS: Microorganism colony attached to the biochar surface

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3.4 Enhanced microbial arsenic transformation and migration with biochar

To precisely describe the transformation and migration of As as well as to illustrate the reasons for As removal from the solution, we measured As species both in aqueous phase and in solid phase. A general trend of As transformation was verified, where As(V) was reduced to As(III) and then converted to organic As. In the blank control, only a negligible portion of As(V) was converted to As(III), and no As was removed (Fig. 4a, b), while in the presence of microorganisms, As(V)aq decreased sharply (Fig. 4c–f). As(III) was the dominant inorganic species in the middle and late stages. In the BC-amended system, a certain amount of As existed in solid phase combined with BC and microbes throughout the whole period (Fig. 4e, f). Since it has been demonstrated that BC presented almost no adsorption capacity to As by our adsorption experiments (Additional file 1: Figure S3b) and many existing studies, it can be ascertained that As was mainly assimilated by microbes (Wu et al. 2020). Owing to the molecular physical similarity to phosphate, As(V) enters microorganisms through phosphate transporters and then participates in metabolism (Mawia et al. 2021). The steady existence of As in solid phase resulted from the microbes’ detoxification (Chen et al. 2020). The rapid transformation of As(V) to As(III) was adverse to microbes, then As(III) was further transformed into methylarsonic acid, which has a weaker intracellular accumulation capacity and arsine, which can escape from the system via volatilization (Li et al. 2021; Savage et al. 2018). The whole process was consistent with microbe detoxification.

Fig. 4
figure 4

The distribution of As(V) and As(III) in aqueous [As(V)aq and As(III)aq] and biochar solid phase [As(V)s and As(III)s] of the co-removal systems along the process. (a), (c) and (e) refer to the low concentration system, while (b), (d) and (f) indicate the high concentration system

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The results  showed that MMAA and DMAA appeared in aqueous solution of the BC-amended systems on day 3 and day 12 (Additional file 1: Table S4). The appearance of methylated As on day 3 indicated that part of the rapidly reduced As(V) had entered the methylation pathway. However, neither MMAA nor DMAA was produced in the No BC systems, and qualitative determination of volatile arsines also verified the conjecture that No BC systems failed in methylation. These phenomena strongly demonstrated that BC promoted As methylation probably because the abundance  of the As(III) methylating gene (AS3MT) was higher in the whole system with the amendment of BC (Yang et al. 2020), and this promoting effect has been widely applied in practical technology (Zhai et al. 2021). The toxicity of most organic As speciations is less than that of As(V), and more less comparable to that of As(III) (Hughes 2002). In future practical application of this process, the system would be airtight, and the released volatile products would be collected and treated together with the harvested biochar carrying the As-laden microorganisms.

3.5 Enriched As detoxification genes in biochar-amended system

In order to resist the As toxicity, bacteria expressed a series of functional genes to achieve As detoxification, which were divided into four categories: As-transforming genes (K03741-arsC, K00537-arsC, K07755-AS3MT), As-migrating genes (K16322-pit, K02040-pstS, K02038-pstA, K02036-pstB, K02037-pstC, K02440-GLPF, K01551-arsA, K03893-arsB, K03325-arsB), As-resisting gene (K11811-arsH), and As-transcriptional repressing gene (K03892-arsR). The transformation of As speciation alongside the biological detoxification with the functional gene in each step is shown in Additional file 1: Figure S8. As(V) firstly enters microorganisms via phosphate transporter (Pit and Pst genes) (Tang et al. 2021), then it is reduced to As(III) by arsC gene with subsequent efflux of As(III) via arsA or arsB genes (Bermanec et al. 2021). Another pathway for As efflux was the methylation of As(III) via AS3MT gene (Hu et al. 2021; Torbøl Pedersen et al. 2020). As-transforming genes were indispensable in the As removal process for detoxification. Meanwhile, As(III) could be absorbed into cell again through GlpF gene to maintain the internal and external balance of As (Tang et al. 2021).

The heatmap of the functional genes in different systems showed that compared to the “No BC” system (307,882 gene units), the BC-amended systems contained more functional genes (354,190 gene units) (Fig. 5). Moreover, the distribution of genes of different functions differed between the liquid environment and biochar surfaces. Firstly, there existed a group of genes, including arsH, K03893-arsB, pit, AS3MT and K00537-arsC genes, which represented the full set of genes required to achieve As(V) detoxification, and the relative abundance of these As(V) detoxification related genes was significantly higher in the aqueous solution (54,589) than on the biochar surface (27,640). This indicated that the complete process of As detoxification mainly occurred in liquid environments. Secondly, another group of genes, consisting of Pst genes (pstA, pstB, pstC and pstS genes), GlpF gene and K03325-arsB gene were dominant on BC surface at early stage, but became dominant in the aqueous solution at late stage (Fig. 5). These genes were related to phosphate transporter responsible for As-adsorption by microorganisms, but did not contain the As-transforming genes. It can be speculated that the dechlorinating microorganisms on the BC surface at early stage adsorbed As into their bodies, and excreted it without transformation. With the progress of dechlorination and proliferation of dechlorination bacteria, these As-adsorption genes released to solution gradually and facilitated the microorganisms to detoxify themselves via transforming the As speciation.

Fig. 5
figure 5

Heatmap of the functional genes. Darker and lighter blue indicate the relative abundance of each gene in different samples depending on Z-score. LC: Low concentration; HC: High concentration; NB: Microbial samples in “No BC” systems; AS: Microorganism suspended in the aqueous solution; BS: Microorganism colony attached to the biochar surface

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3.6 Upregulation of amino acid metabolism for enhanced  As tolerance

General KEGG pathways were classified into 6 categories according to the gene functions (Additional file 1: Figure S9). Three pairs of systems were selected to study the abundance of extracellular metabolites with the comparison of “No BC” and “BC addition”, low and high pollutant concentrations, and early and late process stages. In this system, we only focused on the amino acids with relatively high abundance, significant difference, and apparent regularity, which correlated well with arsenic toxicity. As shown in Fig. 6a, compared with the “No BC” system, a much lower amount of phenylalanine (0.39 times) was detected with BC amendment, while the biosynthesis of phenylalanine was enhanced (Additional file 1: Table S5). This indicated that the presence of BC increased the microbes’ consumption of extracellular phenylalanine, simultaneously upregulated the biosynthesis of phenylalanine for meeting the needs of microbes, which can enhance their As tolerance (Gushgari-Doyle and Alvarez-Cohen 2020). To our surprise, the abundance of serine and lysine tended to increase with high As concentrations (Fig. 6b), while the biosynthesis of them was down-regulated (Additional file 1: Table S5). That meant microbes could not accumulate enough amino acids as an effective approach to resist As toxicity. In other words, the activity of microbes was inhibited by  high As concentrations. In Fig. 6c, both phenylalanine and lysine show the same pattern with analysis of phenylalanine in comparison between HC-BC-LS and HC-NB-LS. That is, both biosynthesis and consumption of these two amino acids were stronger at late stage which indicated more amino acids were required for consistent As damage over time. All of the above conclusions were further confirmed by other comparisons listed in Additional file 1: Table S6 illustrating that increasing amino acids assimilation was also an effective pathway of defense against As toxicity.

Fig. 6
figure 6

Volcano plots showing changes in extracellular metabolite abundance between different comparisons. The horizontal and vertical dotted lines indicate the cutoffs for statistical significance. The colored markers indicate statistically significant metabolites that were present in higher (orange) or lower (blue) abundances. LC: Low concentration; HC: High concentration; ES: Supernatant collected at early stage; LS: Supernatant collected at late stage

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A total of 473 defined metabolites with obvious statistical significance (P < 0.05) were identified among the selected 8 systems. To determine the differential metabolic pathway, metabolite-level changes were imported into MetaboAnalyst 5.0 and analyzed using the metabolic pathway impact analysis module (Fig. 7). Only tryptophan metabolism pathway possessed statistically significant difference (P = 0.023), and there were 7 among 22 differential metabolites matched with this pathway including 5-Hydroxy-l-trytophan, L-Tryptophan, Serotonin, Tryptamine, (Indol-3-yl) acetamide, 5-Hydroxyindoleacetate and Indole-3-acetate. Tryptophan is used in the synthesis of 3-indolepropionic acid, a potent antioxidant used to mitigate DNA damage (Lee et al. 2021). However, both As(III) and As(V) have been reported to promote DNA mutagenesis in bacteria (Zhuang et al. 2021). To verify the correlation of the above two phenomena, we analyzed the genes for mitigating DNA damage involved in genetic information processing (folding, sorting and degradation, replication and repair, transcription, translation) and nucleotide metabolism (Additional file 1: Figure S10). The results showed that these mitigating genes were more concentrated on the BC surface than in the solution. The relative abundances of key replication and repair genes were 1.27 times higher in the presence of BC than in its absence, which was strong evidence for BC promoting the tryptophan metabolism. Therefore, the observed better performance of systems containing biochar can be linked to the enhanced tryptophan metabolism.

Fig. 7
figure 7

Significance analysis of factors impacting metabolic pathways for all cultures. The horizontal dotted line indicate the cutoffs for obvious statistical significance (P < 0.05). Significance ranges from moderately (shallow blue) to highly significant (blue). The size of circles correlates with the number of metabolites identified in the LC–MS analysis for each impacted pathway

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It was noteworthy that in the absence of BC, the content of these DNA damage mitigating genes in the high concentration systems was lower (0.93 times) compared to the low concentration systems (Additional file 1: Figure S10). This implies that the high toxic environment might suppress the ability of microbes to produce DNA damage mitigation genes, and this explains why incomplete dechlorination happened in the HC systems with no BC. As shown in Additional file 1: Figure S11, the functional genes in the high concentration systems with no BC mainly assembled in the second quadrants within a small area. This differed from the high concentration systems with biochar, and low concentration systems, and this phenomenon exactly corresponded to the former inference.

4 Conclusions

This study proposes a technology for simultaneous removal of trichloroethylene (TCE) and arsenic (As) from the co-contaminated groundwater with the addition of biochar into a biodechlorination system. With the assistance of biochar, TCE was completely dechlorinated and finally converted into nontoxic ethene within 12 days even at a relatively high initial concentration (TCE: 30 mg L−1; As(V): 4 mg L−1) mainly via dechlorinating microorganisms (Clostridium and Dehalococcoides). Most of As (50–70%) was removed via separation of biochar colonized by detoxification microbes possessing crucial As-transforming genes (K00537-arsC and K07755-AS3MT) and upregulated amino acid metabolism. Roles of biochar included: (1) adsorbing a part of TCE and the more toxic intermediates such as vinyl chloride, which relieved toxicity stress; (2) facilitating the preferential colonization of dechlorinating microorganisms, while spatially isolating and suppressing the adverse hydrogen-competing microorganisms in the solution; (3) enhancing the microbial tolerance to As(V)/As(III) by upregulating amino acid metabolism and assisting microbes to produce more mitigating DNA damage genes with their self-detoxification ability strengthened. This technology has a great potential to be applied in a real remediation process, and future studies should pay attention to real conditions with coexistence of more contaminants.