Evaluation of degradation behavior over tetracycline hydrochloride by microbial electrochemical technology: Performance, kinetics, and microbial communities

Ecotoxicology and Environmental Safety 188 (2020) 109869

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Evaluation of degradation behavior over tetracycline hydrochloride by microbial electrochemical technology: Performance, kinetics, and microbial communities

T

Xinhong Penga,b,∗, Junrui Caoa, Baolong Xiea, Mengshan Duana, Jianchao Zhaoa a

The Institute of Seawater Desalination and Multipurpose Utilization, Ministry of Natural Resources (MNR), Nankai District, Tianjin, 300192, PR China MOE Key Laboratory of Pollution Processes and Environmental Criteria, Tianjin Key Laboratory of Environmental Remediation and Pollution Control, Nankai University, Tianjin, 300071, PR China

b

A R T I C LE I N FO

A B S T R A C T

Keywords: Tetracycline hydrochloride Microbial electrochemical technology Degradation behavior Bioelectricity generation Microbial community function

Tetracycline hydrochloride (TCH), as a typical antibiotic-pollutant, is desired to enhance its removal from public environment, due to its toxicity and persistence. Microbial electrochemical technology (MET) is a series complex microorganisms-driven processes with characteristics of simultaneous wastewater treatment and electricity generation. The study was presented to evaluate the TCH removal behavior and power generation performance through the co-metabolism under constant glucose with different TCH concentrations using MET. It was found that the TCH removal efficiency arrived at 40% during the first 6 h, when TCH concentrations ranged from 1 to 50 mg/L. It was interesting that TCH degradation rate increased to a maximum of 4.15 × 10−2 h−1 with its concentrations varying from 1 to 20 mg/L, however, the further increase to 50 mg/L in TCH concentration resulted in a reverse 66% reduction. In the meantime, the generated bioelectricity declared a similar fluctuation trend with a maximum power density of 600 mW/m2 under the condition of 20 mg/L TCH co-degradation with glucose. What's more, the TCH inhibition effect fitted well with Haldane's model, indicating that the microbial electrochemical system had a better potency toward TCH toxicity than that reported (EC50 = 2.2 mg/L). Thauera as mainly functional aromatics-degrading bacteria and Bdellovibrio against bacterial pathogens, only existed in the mixed cultures with TCH and glucose, indicating extremely remarkable changes in bacterial community with TCH addition. In summary, a new approach for the anaerobic biodegradation of TCH was explored through cometabolism with glucose using MET. The results should be useful for antibiotics wastewater disposal of containing TCH.

1. Introduction Antibiotics, with characteristics of extraordinary anti-bactericidal abilities, have been widely applied in human health and breeding industry (Cai et al., 2018). In the natural environment, on the one hand, they are excreted into surface water and groundwater via urine or feces as the destiny, which results in the serious aquatic pollution (Wang et al., 2015b). On the other hand, it is inevitable to cause the antibiotic pollution load in soil from farming irrigation, due to the wide use of domestic sewage and livestock wastewater containing antibiotics (Barber et al., 2009; Li et al., 2016; Zhao et al., 2010). The active substances can be washed off from the top soil after rain, and contribute to an increase in the total concentration of antibiotics in sewage and surface water (Kümmerer, 2009). The existence of antibiotics in aquatic

environment would induce the alteration in microbial population, and pose a potential threat to its aquatic ecological process. Thus, we can imagine that the ecosystem function is badly disturbed (Martins et al., 2017; Yi et al., 2019; Zhang et al., 2019). The representative antibiotics (e.g., tetracycline, sulfamethoxazole and lincomycin) and their metabolites can be universally monitored in the northern China (Hu et al., 2010). Among those, tetracycline is a kind of chemicals with low solubility due to the occurrence of free tertiary amine structure. Once it is neutralized by hydrochloric acid, it would change into water-soluble and alkyl substituted ammonium salt in the form of tetracycline hydrochloride (TCH, C22H24N2O8·HCl) with a stable four-ring structure (Cai et al., 2018; Leclercq et al., 2016; Qian et al., 2016). Here, TCH belongs to a family of broad-spectrum antibiotics that inhibit protein synthesis in bacteria by blocking the binding of aminoacyl-tRNA to the

∗ Corresponding author. The Institute of Seawater Desalination and Multipurpose Utilization, Ministry of Natural Resources (MNR), Nankai District, Tianjin, 300192, PR China. E-mail address: [email protected] (X. Peng).

https://doi.org/10.1016/j.ecoenv.2019.109869 Received 23 July 2019; Received in revised form 22 October 2019; Accepted 23 October 2019 0147-6513/ © 2019 Elsevier Inc. All rights reserved.

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described previously (Peng et al., 2012) until the output voltage was ≥300 mV. The electrolyte was completely refreshed when the cell voltage was lower than 50 mV. It was considered to be suitable for TCH degradation behavior tests when the similar cell voltage output was produced over three consecutive cycles. During the process, TCH contents were varied in compliance with 1, 2, 10, 20, 30 and 50 mg/L with constant glucose level to describe the effects of TCH initial concentration on its removal and power generation. All chemical reagents were used in analytical grade without further purification.

bacterial ribosome, and have no antifungal activity due to the un-ability to cross the fungal cell wall (Basile et al., 1998). More and more literatures have been reported that TCH is too recalcitrant to be efficiently converted by traditional biological treatment processes (Xiong et al., 2017) or physicochemical methods including adsorption, precipitation, hydrolysis and complexation (Liu et al., 2014; Sassman and Lee, 2005; Wang et al., 2017; Yang et al., 2011). Therefore, there is an urgent need to seek a new and high-efficient approach to destruct its complex structure and remove its biotoxicity. In recent years, the use of microbial electrochemical technology (MET) to directly convert chemical energy to electricity with simultaneous wastewater treatment has attracted extensive attention due to its low-cost, short treatment time, and high efficiency by the process of biological metabolism coupling with electrochemistry (Du et al., 2017; Peng et al., 2018; Wang and Ren, 2013). It has been reported that the electroactive biofilms formation can be enhanced by trace tobramycin as an agonist in microbial electrochemical system (Zhou et al., 2017). Based on the co-metabolism principle of parent contaminant and substrate, the research has exhibited that the application of MET can effectively eliminate the biotoxicity of sulfadiazine (Wang et al., 2018). Zhou et al. have not only distributed microbial fuel cell (MFC) to remove antibiotics (aureomycin, roxithromycin and norfloxacin) with the efficiency of 100%, but promoted the removal of NH3–N and TP by the addition of antibiotics (Zhou et al., 2018). It is encouraging with the exceptional benefits in antibiotics pollution abatement and bioelectricity production by MET. Excitingly, Yan et al. suggested that MET is beneficial to the reduction of antibiotic resistance genes (ARGs) compared with traditional biological treatments (Yan et al., 2019). However, there are few studies technically aiming at TCH de-toxicity by the approach of MET. Li et al. have found that the removal of 30 mg/L TCH was 90% by FeOOH/TiO2 granular activated carbon as expanded cathode in low-cost MBR/MFC coupled system (Li et al., 2017). Interestingly, how the initial concentration of TCH affect its removal rate is not yet clear, moreover, the related mechanism and the degradation pathway on TCH removal are poorly understood. Consequently, this study was designed to explore the fate and behavior of TCH-destruction with the synergistic catalysis of microorganism and electrochemistry using MET. Hence, the objectives of this study included: i) characterizations of TCH- removal and its corresponding bioelectricity performance through anaerobic co-metabolism by the process of MET; ii) kinetics and mechanism of TCH-degradation, including the degradation efficiency and byproducts identification; iii) microorganism community succession contributing to TCH removal. Overall, this study was devoted to provide a different method for TCH degradation and analysis of the efficiency of the whole process by MET.

2.2. Analytical methods Quantitative and qualitative analysis for TCH TCH concentration was quantified by the high-performance liquid chromatography (HPLC, Agilent 1200 series, USA) equipped with a diode array detector (DAD) and a Zorbax Eclipse XDB-C18 reversed-phase column (4.6 mm × 150 mm × 5 μm). 60% acetonitrile and 40% deionized water (containing 0.1% formic acid) were used as the mobile phase, and the flow rate, detection wavelength, column temperature was 1.0 mL/ min, 273 nm and 30 °C, respectively. In order to qualify the intermediate products from TCH decomposition, Waters ACQUITY UPLC H-Class equipment coupled to tandem mass spectrometry (Waters XEVO ® TQD) was adopted for the analysis equipping with an electrospray ionization source (ESI). The MS analysis was performed in full scan mode of 10–2000 Da with a spray voltage of 3.5 kV, and a capillary temperature of 150 °C. Equipment control and data acquisition were performed with Masslynx v. 4.1 software (Waters, USA). Thus, the molecular structure for its transformation products were tentatively proposed by the detection of predicted mass (Jia et al., 2009; Leng et al., 2016). Before the LC-ESI-MS/MS detection, the sample was filtered by 0.22 μm microporous membrane, and then further loaded into solid phase extraction (SPE) at a flow rate of 1 mL/min to purify with Cleanert PEP (500 mg, 6 mL) which had been activated with 3 mL methanol and 3 mL distilled water in sequence. Subsequently, the SPE column was eluted with 3 mL distilled water followed by 2 × 3 mL methanol solution, and then the eluate was concentrated under a nitrogen stream in 45 °C water bath. Electricity performance analysis The output voltage (V) across resistor was recorded every 30 min using a date acquisition system (PISO-813, ICP DAS Co. Ltd, China) through a personal computer. The current (i) was calculated according to the Ohm's Law (i = V/R), here, R was the external resistance, and the power density (P) was evaluated as P = V2/R. The polarization and power density curves, normalized on the cathode surface area, were generated by the variation of R from 1000 to 50 Ω with a 20-min time interval. Genetic characteristics analysis Bacterial samples were denoted as S0, S1 and S2, which were collected from the glucose/TCH co-cultures, the inoculum culture and pure glucose, respectively. Genome DNA was extracted by the method of CTAB/SDS, and confirmed its purity by 1% agarose gel electrophoresis with diluting to 1 ng/μL using sterile water. PCR-amplification was carried out by 16S rRNA gene fragments using specific primers (GTGCCAGCMGCCGCGGTAA, GGACTACHVGGGTWTCTAAT, 16S V4: 515F-806R) with Phusion®HighFidelity PCR Master Mix (New England Biolabs). The mixture of PCR products was purified under equidensity ratios with QiagenGel Extraction Kit (Qiagen, Germany) followed being detected by 2% agarose gel electrophoresis. The library was made up by TruSeq®DNA PCR-Free Sample Preparation Kit (Illumina, USA) in the light of manufacturer's recommendations, and quantified by the [email protected] Fluorometer (Thermo Scientific) and Agilent Bioanalyzer 2100 system, and then sequenced on an Illumina HiSeq2500 PE250 platform.

2. Materials and methods 2.1. Experimental configuration All the single-chamber air-cathode configurations, made of the polymethyl methacrylate, were used as the MET reactors (8 cm long by 5 cm diameter; 150 mL effective volume) running at 30 ± 0.5 °C temperature-controlled biochemical incubator (Taisite Instrument Co. Ltd, Tianjin, China) in parallel with batch mode under 1 kΩ except as noted otherwise. The carbon-mesh brush electrode (3 cm in diameter and 3 cm long) serving as MET anode, was subjected to a 24 h acetone soak followed by a 30 min heat treatment at 450 °C in a muffle furnace, and then deployed vertically with the activated carbon air-cathode to avoid the power overshoot (Peng et al., 2013) by an anti-corrosion titanium wire connection. The inoculum from the secondary effluent of Tianjin municipal domestic sewage treatment plant without antibiotics disposal, was fed with 1 g/L glucose as the carbon source under the ratio of 1:4 (v/v) to nutrient mediums: 50 mM phosphate buffer solution (PBS, g/L, Na2HPO4 4.576, NaH2PO4 2.132, NH4Cl 0.31, KCl 0.13) with the pH of 7.0, trace minerals (12.5 mL/L) and vitamins (5 mL/L) as

2.3. Computation methods TCH removal efficiency (RTCH) for this study was calculated as 2

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shown in equation (1). Linear regression models with a 95% confidence interval were conducted to establish significance levels (p < 0.05). equation (2) was employed to fit the reaction rate constant (k) using first-order kinetics with the TCH concentration variation, correspondingly, the half-life (τ) of TCH degradation could be determined according to equation (3).

RTCH = (C0 − Ct )/ C0.

(1)

lnCt = −kt + lnC0

(2)

τ = ln2/ k .

(3)

Table 1 Kinetic equation and parameter of TCH degradation by MET.

Vmax C Ks + C +

C2

Ki

Fitting equation

k (h−1)

τ (h)

1 2 10 20 30 50

y y y y y y

0.0129 0.0329 0.0343 0.0415 0.0162 0.0141

53.3 21.1 20.2 16.7 42.7 49.2

= = = = = =

0.0129x + 0.420 0.0329X + 0.007 0.0343x + 0.106 0.0415x + 0.380 0.0162x + 0.398 0.0141x + 0.250

increased by 42% to ca. 78 h to attain the removal efficiency of 80%. In other words, TCH degradation reaction rates were increased with the increase in TCH initial concentration ranging from 1 to 20 mg/L, nevertheless, the further increase of TCH concentration to 50 mg/L resulted in a quick decrease for the TCH treatment performance, indicating that TCH degradation behavior could be facilitated with the appropriate concentration using MET, otherwise, the microbial electrocatalytic activity would be inhibited over the optimum of TCH concentration. Therefore, in order to achieve a superior degradation performance, it is recommended to apply a suitable concentration for TCH wastewater disposal. Kinetics of TCH degradation It is kinetics of a bioreaction to be the most obvious feature (Chen et al., 2011). Therefore, batch assays were carried out to determine the kinetics of TCH degradation. The experimental data of Table 1 described good linear fitting profile using the first-order model with the corresponding R2 value (> 0.98). It was 3.2 folds for TCH degradation rate constant (k) from 1.29 × 10−2 to 4.15 × 10−2 h−1 with its initial concentrations ranging from 1 to 20 mg/L. Further increase of TCH concentration to 50 mg/L resulted in an evident decrease of 66% for k to 1.41 × 10−2 h−1. Correspondingly, the half-times (τ) of pollutant degradation reaction were firstly decreased from 53.3 h for 1 mg/L TCH to 16.7 h for 20 mg/L TCH, and then reversely increased to 49.2 h for 50 mg/L TCH, suggesting that TCH degradation behavior could be promoted by the MET co-metabolic process with its initial concentration no more than 20 mg/L, but the increasing concentration would inhibit the catalytical ability of microbial electrochemical system for TCH to bring about the lower degradation efficiency. In other words, the TCH removal performance is limited with the increase in its initial concentrations by MET. Take the substrate inhibition into account, Haldane model was adopted to match the experimental data with TCH degradation kinetics factors in this study. It was well stated the dynamic inhibiting characteristics for TCH removal under different initial concentrations with the correlation coefficient (R2) of 0.99. Noticeably, the specific degradation rate was low with 30–50 mg/L TCH, possibly due to the inhibitory effect of high concentration for TCH (> 20 mg/L) to microbial electrochemical activity. This was in accordance with the fluctuation of degradation efficiency, here, we could see that it was a bit higher for 30 mg/L TCH than that of 50 mg/L TCH (Fig. 1). The Vmax , Ks , and Ki were 15.1 × 10−2 mg/L/h, 5.09 and 8.48 mg/L, respectively. Thus, the model was:

where “t” was the reaction time, “lnCt ” and “lnC0 ” were natural logarithms of target pollutant concentration (Ct and C0 ) at “t” and 0 time, respectively. The influence of initial pollutant concentration on degradation rate was analyzed using Haldane's inhibitory model as follows:

V=

Concentration (mg/L)

(4)

where C is the TCH concentration (mg/L), Vmax is the maximum specific degradation rate (mg/L/h), Ks is the half-saturation constant (mg/L), and Ki is the inhibition constant (mg/L). The software Origin (version 9.5) was adopted for the predictive analysis. 3. Results and discussion 3.1. Degradation behavior of TCH removal Effect of initial TCH concentrations In the experiment, the level of glucose was constant with different TCH concentrations as the co-substrates. The concentration conversion of TCH was plotted against the elapsed time in Fig. 1 to discuss the effect of initial concentrations. It showed that a remarkable feature was the sharp reduction of the TCH content with its removal rate of 40% in the first 6 h for every TCH initial concentration, possibly due to the rapid oxidation of the parent pollutant with the help of biocatalysis. However, the following removal rate of TCH was noticeably influenced with the microbial electrochemical reaction proceeding (Fig. 1). As it displayed, in the co-culture with TCH concentration of 2 mg/L, it was 55 h for the time required to arrive at an 80% removal, whereas in that of 20 mg/L for TCH concentration, the required degradation time decreased considerably by 49% to 28 h to arrive the same removal behavior. Unfortunately, with the further increase in TCH initial content to 50 mg/L, the demanding time reversely

15.1 × 10−2 C

V=

5.09 + C +

C2

.

8.48

According to the above-mentioned model, an optimal concentration (Sopt ) and degradation rate (Vopt ) could be respectively calculated as follows:

Sopt =

Ki × Ks .

Vopt = Vmax /(1 + 2 Ks / Ki ). Therefore, in the study, the Vopt value for TCH degradation was 5.87 × 10−2 mg/L/h, and the Sopt is 6.56 mg/L with a higher tolerance to biotoxicity than that reported (EC50 = 2.2 mg/L) (Halling-Sørensen, 2001), which was probably resulted from the abundant biological

Fig. 1. Effects of initial TCH concentration on its degradation. 3

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Fig. 2. The total ion chromatograms (TIC) diagram and mass spectrogram of with different retention time.

degradation pathway of TCH using MET was proposed (Fig. 3) in the presence of glucose as the co-culture. The ion products at m/z 445 were designated to be the TCH molecule (Wang et al., 2015a). The intermediate with m/z 416 was generated due to the loss of the N-methyl group by the demethylation of the dimethylamino group at position C4 by biotransformation. With the further removal of the carbonyl and amine groups, the compound m/z 370 was formed, which was in accordance with TCH oxidation by potassium ferrate (Jiao et al., 2008). Thus, it was understood that TCH transformation products had lower toxicity than the parent compound. With the TCH's 4-ring molecule structure being destroyed, some small molecular intermediates with m/ z values of 225, 245 and 297 were produced through the open-ring

populations using MET. Thus, it was inferred that TCH at the concentration of 20 mg/L actually acted as both an antibiotic to the bacteria cells and a protein inactivation, resulting in the loss of living organisms (Xiong et al., 2017). This was similar with the decolorization of Reactive Black 5 as observed in the literature (Wang et al., 2013). Degradation Pathway for TCH Removal It is not easy to dispose for TCH, due to the stable four ring structure. In light of environmental perspective, it is more desirable for microbial transformation of TCH than physicochemical processes, because antibiotic toxicity can be reduced by the biological degradation behavior. The possible intermediate products or by-products of bio-decomposition for TCH were denoted from the UPLC-ESI-MS/MS spectra (Fig. 2), and the possible 4

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Fig. 3. Proposed biodegradation pathway of TCH.

reaction and cleavage of the central carbon. It was assumed that the intimate coupling of electrochemical catalysis and biodegradation could enhance the TCH dissociation. The similar result was also observed during p-chloronitrobenzene degradation by MET (Peng et al., 2018).

3.2. Bioelectricity performance with TCH addition Bioelectricity is one of the mainly theoretical approaches for performance evaluation in microbial electrochemical system. As soon as the system went into the steady running state, the power generation was checked out for the optimum performance with a maximum power density (MPD) of 524 mW/m2, which was reached at a current density of 1.58 A/m2 without TCH containing in the medium except for the glucose (Fig. 4a). At this moment, the external resistance was 300 Ω, and based on this, the subsequent bioelectricity performance was conducted under different TCH contents. It revealed that there was a slighty positive increase of 15% to 600 mW/m2 in the MPD when the TCH addition ranged from 1 to 20 mg/L. With the further increases to 50 mg/L in TCH concentration, a negative reduction of ca. 9% in MPD was obtained to achieve 548 mW/m2 at the current density of 1.614 A/ m2 (Fig. 4b). It was noted that the bioelectricity generation was enhanced when TCH concent increased to 20 mg/L, however, the further increment in TCH content didn't result in a correspondent improvement on power performance. This was in accordance with the result from the TCH removal behavior. As depicted above, the exoelectrogens had a better tolerance toward TCH biotoxicity for MET than reported. Thus, it was enough for eletrons to oxide the co-substrates of TCH and glucose under the condition of lower TCH concentration (≤20 mg/L), which brought about the power performance enhancement by TCH addition. While the electron consumption between the co-cultures was competitive with the TCH content increasing, the TCH toxicity to electroactive biofilms became predominant, resulting in the inhibition to bioelectricity. The similar results have been reported in the early studies. For example, the MPDs arrived at 45.4 and 51.2 mW/m2 with 1000 mg/L glucose, a mixture of 1000 mg/L glucose and 250 mg/L indole as the MET substrates, respectively (Luo et al., 2010). From the point of practical application, glucose, as the co-substrate, was not advisable in virtue of the increase of operating cost. However, to put that into context, the large amounts of biodegradable organic matter in wastewater can be exploited as electron donors to improve the pollutant degradation effectiveness and power generation performance.

Fig. 4. Polarization curve for optimum working condition (a) and power performance based on different TCH concentration (b).

5

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abundance of 23.8%, 26.1% and 17.3%, respectively. After cultured in the presence of pure glucose (S2), the relative abundance (23.5%) of Bacteroidetes significantly increased by 35.8%, while there was almost no change for Proteobacteria (25.9%) and Firmicutes (25.7%) in the system of S2. On further shift to mixed substrates with the additive TCH, the microbial community displayed that there was apparently positive increase of 3.3 folds (78.5%) for Proteobacteria and reverse decrease of 95% (1.3%) for Firmicutes. Besides, the presence of TCH led to a relative abundance of 0.43% and 0.21% for Nitrospirae and Chloroflexi in the sample of S0, in contrast, there were little in the systems of S1 and S2, indicating that the TCH addition could boost the oriented enrichment of the leading predominant bacteria. Fig. 5b elucidated that Thiovirga, was the dominant bacteria with the most abundant genus (44.6%) in the co-substrates of TCH and glucose (S0) in comparison with S1 (0.2%) and S2 (0.03%), which was closely related with wastewater treatment and resource recovery processes (Kappell et al., 2018). While, Pseudomonas obviously declined from 13.7% (S2) to 0.38% (S0) in virtue of TCH addition under the electric-stimulation. Thauera, as the second dominant genus, displayed a variation trend of high - low - low (4.2%–0% - 0%) in relative content of S0, S1 and S2, respectively. It has been reported that Thauera not only exists in phenol-contaminated soils and sewage systems, but is helpful of effective degradation of aromatic hydrocarbons (Wang et al., 2018; Zhao et al., 2018). What's more, Thauera has a uniquely good adaptation of antibiotic pressure (Mao et al., 2010). In addition, Bdellovibrio, detected with the abundance of 2.4% in S0 (0% in S1 and S2), has a promising potential characteristic against bacterial pathogens

Table 2 Comparison of community diversity and richness between co-cultures of glucose and TCH (S0), inoculum culture (S1) and pure glucose culture (S2). Sample name

Chaol

ACE

Shannon

Simpson

S0 S1 S2

649.778 246.833 266.240

656.286 244.098 263.704

5.047 3.867 4.419

0.823 0.854 0.899

3.3. Microbial community structure and function In order to better explain the effect of microbial community on TCH degradation, alpha-diversity indices was calculated to describe the microbial species richness and diversity as shown in Table 2. The Chao 1 and ACE indices present the species richness, while the Shannon and Simpson indices estimate the microbial community diversity (Zhang et al., 2018). In the study, it was significantly different for bioelectrode (S0) existing with highest richness and diversity compared to those of the other two samples, indicating that the biological diversity could be improved by the additive domestication of TCH under the same electrical stimulation. It is also reported that the high bacterial biodiversity is conducive to increase degradation performance of hydrocarbons (Dell'Anno et al., 2012). Fig. 5a showed the sequencing results to phylum-level of microbial community structure from the initial inoculated source to the other two samples. Proteobacteria, Firmicutes, Bacteroidetes were the three dominantly functional phyla for the inoculum bacteria (S1) with the relative

Fig. 5. Microbial community identified at the phylum level (a) and genus level (b) between co-cultures of glucose and TCH (S0), inoculum culture (S1) and pure glucose culture (S2). 6

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(Shatzkes et al., 2015). Thus, it can be inferred that Thauera and Bdellovibrio are conducive to accelerate the TCH degradation with their enrichment over time.

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4. Conclusions This work demonstrates that TCH degradation in the environment is possible through MET with higher microbial reactivity. With the elevated TCH concentration under the condition of anaerobic co-metabolism, both TCH degradation kinetic rate and power generation performance increase first and then decrease under the optimal content of 20 mg/L TCH. Thauera as the mainly functional bacteria and Bdellovibrio against bacterial pathogens, are only disclosed in the mixed co-substrates. Thus, it provides the proof-of-concept of the application of MET to wastewater treatment for other aromatic antibiotics. However, continued study of TCH degradation in MET is still needed to clear the alterations in microbial community and the stability in longterm operation, and thus, provide the accurate assessment of general applicability for MET in TCH removal. Acknowledgements This work was supported by National Key Research and Development Program of China (No. 2017YFC0403502), National Natural Science Foundation of China (No. 51409052) and the Central Public Interest Scientific Institution Basal Research Fund (No. KJBYWF-2019-T06). We also thank the Shanghai Tongji Gao Tingyao Environmental Science & Technology Development Foundation (STGEF). References Barber, L.B., et al., 2009. Fate of sulfamethoxazole, 4-nonylphenol, and 17 beta-estradiol in groundwater contaminated by wastewater treatment plant effluent. Environ. Sci. Technol. 43, 4843–4850. Basile, A., et al., 1998. Antibiotic effects of Lunularia cruciata (Bryophyta) extract. Pharm. Biol. 36, 25–28. Cai, M., et al., 2018. Systematic characterization and proposed pathway of tetracycline degradation in solid waste treatment by Hermetia illucens with intestinal microbiota. Environ. Pollut. 242, 634–642. Chen, G., et al., 2011. A batch decolorization and kinetic study of Reactive Black 5 by a bacterial strain Enterobacter sp GY-1. Int. Biodeterior. Biodegrad. 65, 790–796. Dell'Anno, A., et al., 2012. High bacterial biodiversity increases degradation performance of hydrocarbons during bioremediation of contaminated harbor marine sediments. Environ. Pollut. 167, 85–92. Du, Q., et al., 2017. Protection of electroactive biofilm from extreme acid shock by polydopamine encapsulation. Environ. Sci. Technol. Lett. 4, 345–349. Halling-Sørensen, B., 2001. Inhibition of aerobic growth and nitrification of bacteria in sewage sludge by antibacterial agents. Arch. Environ. Contam. Toxicol. 40, 451–460. Hu, X., et al., 2010. Occurrence and source analysis of typical veterinary antibiotics in manure, soil, vegetables and groundwater from organic vegetable bases, northern China. Environ. Pollut. 158, 2992–2998. Jia, A., et al., 2009. Simultaneous determination of tetracyclines and their degradation products in environmental waters by liquid chromatography–electrospray tandem mass spectrometry. J. Chromatogr. A 1216, 4655–4662. Jiao, S., et al., 2008. Aqueous photolysis of tetracycline and toxicity of photolytic products to luminescent bacteria. Chemosphere 73, 377–382. Kappell, A.D., et al., 2018. Removal of antibiotic resistance genes in an anaerobic membrane bioreactor treating primary clarifier effluent at 20°C. Environ. Sci.: Water Research & Technology 4, 1783–1793. Kümmerer, K., 2009. Antibiotics in the aquatic environment–a review–part I. Chemosphere 75, 417–434. Leclercq, S.O., et al., 2016. A multiplayer game: species of Clostridium, Acinetobacter, and Pseudomonas are responsible for the persistence of antibiotic resistance genes in

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