The co-effect of ampicillin and multi-walled carbon nanotubes on activated sludge in sequencing batch reactors: microbial status, microbial community structure and ARGs propagation

Abstract

The abuse of antibiotics and the release of nanoparticles (NPs) into the environment are of grave concern; however, their co-effect on the environment is not yet clear. In this study, the response of activated sludge exposed to ampicillin (AMP) and multi-walled carbon nanotubes (MWCNTs) was investigated in a sequencing batch reactor (SBR). The results showed that mixed liquor volatile suspended solid (MLVSS) values exposed to AMP and MWCNTs decreased to 83.48–90.53%, but the specific oxygen utilization rate (SOUR) was even higher with respect to the control for 24 h exposure. Compared to individual AMP, the co-occurrences of AMP and MWCNTs showed lower cytotoxicity to activated sludge, supported by a lower membrane permeability and higher cell metabolic viability. The microbial community structure changed when exposed to AMP and MWCNTs. Firmicutes phylum emerged in the top 20 phyla under the co-effect of AMP and MWCNTs, but not for AMP alone. It is worth noting that the relative abundances of antibiotic resistance genes (ARGs) under the co-effect of AMP and MWCNTs were significantly increased, especially bla_DHA. These results provide useful information for understanding the risk of the co-occurrences of antibiotics and CNTs, and can aid the sustainable and widespread application of novel carbon materials.

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Environmental significance

Both carbon nanotubes and antibiotics could be discharged into wastewater treatment plants. However, their co-effect on the environment is not yet clear. Consequently, this study investigated the response of activated sludge exposed to ampicillin and multi-walled carbon nanotubes in a sequencing batch reactor. The co-occurrences of ampicillin and multi-walled carbon nanotubes showed a lower cytotoxicity to activated sludge compared to the individual materials. However, the relative abundance of antibiotic resistance genes increased after co-exposure to ampicillin and multi-walled carbon nanotubes. These results provided new insight into the potential risk of the co-occurrences of antibiotics and carbon nanotubes.

1. Introduction

Carbon nanotubes (CNTs) are recognized as the “king of nanomaterials” due to their excellent mechanical, electrical and chemical properties.¹ However, the potential ecological and health risks of CNTs have raised extensive concerns along with their widespread applications. CNTs are reported to show some toxic effects on fish, invertebrates, algae and microorganisms in the aquatic environment.^2–4 For example, CNTs could produce a higher microbial oxidative stress response that may damage the integrity of cell membranes.⁵ This results in the change in the microbial community structure and a decrease in the denitrification efficiency of microorganism.^6,7

In recent years, it has been observed that carbon nanomaterials could interact with coexisting pollutants and could change their mobility, toxicity and bioavailability.⁸ CNTs into riverine sediment were found to significantly increase the adsorption amount of 2,4-dichlorophenol in sediment from 0.541 to 1.44 mg g⁻¹, and slightly inhibited the dehydrogenase activity of microorganisms in the sediment.⁹ Other studies showed that the presence of CNTs could protect microorganisms from the cytotoxicity of phenol in wastewater but reduced the bacterial diversity in sequencing batch reactors.¹⁰ Also, it was found that carbon nanomaterials could lead to a significant increment in the intracellular iron concentration in anaerobic wastewater treatment.¹¹ Although many valuable results have been obtained, efforts are still required to advance the understanding of the environmental risks of CNTs with coexisting pollutants.

Antibiotics residues have become a critical environmental and public health issue globally because they could result in the propagation of antibiotic resistant bacteria (ARB) and antibiotic resistance genes (ARGs).^12–14 Recently, it was recognized that the co-occurrences of antibiotics and engineering nanoparticles (ENPs) present synergistic effects on microorganisms.^15–17 Deng et al. found that tetracycline can bind silver nanoparticles to form a tetracycline-nano Ag complex, which can present a synergistic growth inhibition against Salmonella bacteria through strong interaction with the cells and a temporal high concentration of Ag⁺ near the bacteria cell wall.¹⁸ Moreover, it has been found that nano-ZnO and nano-Al₂O₃ could promote the horizontal transfer of ARGs within or across genera in pure cultures or indigenous aquatic microbiota systems by damaging cell membranes.^19–21

According to material flow analysis, both CNTs and antibiotics can be discharged into wastewater and may eventually enter wastewater treatment plants (WWTPs).^22,23 Theoretically, the co-occurrences of CNTs and antibiotics are bound to affect the status of activated sludge in WWTPs. This would pose an environmental risk and even a dissemination risk of ARB and ARGs. However, to the best of our knowledge, there are few reports about this issue in the literature.

Therefore, multi-walled carbon nanotubes (MWCNTs) and ampicillin (AMP), which is widely used in aquaculture, husbandry, and health care due to its low cost and high performance,²⁴ were chosen in this study to explore the co-effect of CNTs and antibiotics on activated sludge in a sequencing batch reactor (SBR). The operating state of the SBR was investigated through analysing mixed liquor volatile suspended solids and the specific oxygen utilization rates. The physiological status of the microorganisms was assessed by membrane permeabilization, cell metabolic viability and cell death mode. Furthermore, the microbial community structure and resistance genes were identified by high-throughput sequencing and fluorescence quantitative PCR. Based on our findings, this study can serve as a reference for a better understanding of the environmental risks of carbon nanomaterials.

2. Materials and methods

2.1 Multi-walled carbon nanotubes stock suspension

In this study, the MWCNTs were purchased from Hengqiu Nano Technology Co. (Suzhou, China). The MWCNTs were more than 95% pure by dry weight, had specific surface areas greater than 874.93 m² g⁻¹, inner diameters of 3–5 nm, outer diameters of 8–15 nm and were 3–12 μm in length. The MWCNTs stock suspension was prepared by dispersing MWCNTs in deionized water for 30 min in an ultrasonic bath followed by 60 min in an ultrasonic cell disruptor (Ningbo Scientz Biotechnology Co., China). The suspension was sonicated again in the ultrasonic bath for 30 min before use to ensure a homogeneous mixture was obtained.²⁵ The size distribution and morphology of the MWCNTs in the stock suspension (0.1 mg L⁻¹) were analysed by dynamic light scattering (DLS) (Malvern Zetasizer Nano ZS90, Malvern Instruments, UK) and transmission electron microscopy (TEM, Tecnai G2 F20, USA), the details of which are depicted in ESI† S1. MWCNTs are thought to be a good potential adsorbent, so an adsorption experiment, described in ESI† S2, was performed to verify their adsorption amount of AMP.

2.2 Exposure experiments

In this study, six SBRs were operated to investigate the co-effect of MWCNTs and AMP. A schematic diagram of them is shown in Fig. S1.† Each bioreactor was a cylinder with an internal diameter of 25 cm and an effective working volume of 10 L. The seed sludge of the SBR was taken from the aeration tank of WWTPs in Qingdao, Shandong. The mixed liquid suspended solids (MLSS) concentration in each reactor was 3500 ± 100 mg L⁻¹ in the initial operation period. The synthetic wastewater was prepared according to ref. 5, 10 and 11 to simulate domestic sewage, and its preparation method and composition are described in ESI† S3 and Table S1. The operating parameters of SBR are described in Table S2.† The whole experimental period was 15 days, which included five days of domestication and ten days of short-term exposure experimentation.

In order to avoid the interaction between the MWCNTs and AMP before entering the bioreactor, the MWCNTs stock suspension was directly fed into the bioreactor in one go as a mixture of sludge after domestication, while AMP was fed as a contaminant in the synthetic wastewater. Although the concentrations of MWCNTs in the wastewater as a contaminant was predicted to be between 6.6 ng L⁻¹ and 18 ng L⁻¹ according to their fade models,²⁶ MWCNTs could accumulate in the sludge due to their high hydrophobic nature.²⁷ Considering the significantly biotoxicity of high doses of MWCNTs in soil and the aquatic environment,³ the MWCNTs concentration in SBR was set from 20 mg L⁻¹ to 100 mg L⁻¹ in this study. The concentration of AMP in the synthetic wastewater used was 50 μg L⁻¹ based on its content of 16–150 μg L⁻¹ in the actual wastewater.²⁸ The concentrations of AMP and MWCNTs, beside the labels of each bioreactors, are shown in Table 1.

Table 1 Concentration of AMP and the total mass of MWCNTs

No.	Name of reactor	Total mass of MWCNTs (mg)	Concentration of MWCNTs (mg L^−1)	Concentration of AMP (μg L^−1)
1	R CP	0	0	0
2	R CNT	1000	100	0
3	R AMP	0	0	50
4	R D20	200	20	50
5	R D50	500	50	50
6	R D100	1000	100	50

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In order to avoid the fluctuation error of SBR, the whole exposure experiments were repeated three times. During each exposure experiment, mixed liquor volatile suspended solid (MLVSS) and the specific oxygen utilization rate (SOUR) of each SBR were measured according to the standard for the sludge stabilization treatment of municipal wastewater treatment plants (CJ/T510-2017). Total organic carbon (TOC) and the concentration of AMP in the effluent of each bioreactor were analysed using the TOC-L analyser (Shimadzu Co., Kyoto, Japan) and high-performance liquid chromatography (HPLC) (Thermo Fisher Scientific, USA), respectively. Sludge samples were homogeneously harvested to be evaluated from each bioreactor at 24 h and at the end of the exposure experiments (240 h).

2.3 Physiological status of the microorganisms

The physiological status of the microorganisms in activated sludge was assessed by various staining procedures, as shown in Fig. 1. The microorganisms in the activated sludge were prepared as single-cells through dilution and sonication according to a previous method.²⁹ A PI staining (Invitrogen, Carlsbad, CA, USA) and Live/Dead cell kit (Calcein-AM and PI, Invitrogen, Carlsbad, CA, USA) were used to evaluate the content of the intact cells in the sludge samples based on the membrane integrity and membrane permeability of the microorganisms.^30,31 The death modes of the microorganisms, including apoptosis and necrosis, were analysed using an Annexin V-FITC/PI apoptosis kit (Invitrogen, Carlsbad, CA, USA) according to the manufacturer's instructions.^32,33 The Alamar Blue assay (AB, China) was used to visually estimate microbial activity according to cellular proliferation. The fluorescence intensity of the PI staining, Live/Dead cell kit and Annexin V-FITC/PI apoptosis kit were tested by Image Stream^X Mark II flow cytometry (Amnis, USA), and the Spark multifunctional microplate reader (Tecan, Austria) was used to measure the fluorescence intensity of the Alamar Blue assay. Details of the extraction and analysis process are available in ESI† S4–S7.

Fig. 1 Physiological status of microorganisms and the test criteria for identification.

2.4 Microbial community and ARGs analysis

Sludge samples harvested at the end of the exposure experiments (240 h) were immediately stored at −20 °C until their DNA was extracted (ESI† S8). 16S-rDNA gene high-throughput sequencing of the activated sludge samples was performed at the Personalbio Genomics Institute (Shanghai, China) using the Illumina Hiseq platform. ARGs, as well as mobile genetic elements (intI1 and intI2), were quantified by fluorescence quantitative PCR (TIB 8600, China) at Sunny Biotechnology Co., Ltd (Shanghai, China). All the information on the primers is given in Table S3.†^34–37 Also, the details of the microbial community and ARGs analysis are available in ESI† S9.

2.5 Other analytical procedures

Extracellular polymeric substances (EPSs) from the activated sludge samples were extracted by a heat method according to the previous method (ESI† S10).³⁸ The content of polysaccharides in EPS were obtained by anthrone-sulfate colorimetry.³⁹

The degree of intracellular reactive oxygen species (ROS) accumulation (including O²⁻, H₂O₂, HO₂˙ and ˙OH) in the activated sludge was determined using a fluorescence probe 2′, 7′-dichlorofluoresceindiacetate (DCFH-DA; Sigma-Aldrich, Saint Louis, MO, USA), according to the manufacturer's instructions.^40,41 The details are in ESI† S11.

2.6 Statistical analysis

Each test in every exposure experiment was conducted in at least triplicate, and an independent sample t-test using SPSS 23.0 (IBM-SPSS, Chicago, IL, USA) by the Personalbio Genomics Institute (Shanghai, China) was used to determine whether there was statistical evidence that the set of test values were significantly different for the evaluation of the microbial community structures. A value of P < 0.05 was noted to be significant, and a value of P < 0.01 was noted to be very significant.

3. Results and discussion

3.1 MWCNTs characteristics in the SBR

The toxicity of MWCNTs largely depends on their dispersed state (such as aggregate diameter and morphology) in an aquatic environment.^15,42 Herein, the preliminary characterizations of the MWCNTs were analysed as they were dispersed in the synthetic wastewater and with the activated sludge in the SBR (Fig. S2–S4†). It was found that the aggregate diameter of MWCNTs fell in the range of 10 to 100 nm as they were added into the wastewater samples in the first 10 min after ultrasound treatment, and then increased to the range of 40 to 200 nm after 240 h. This agreed with the fact that MWCNTs gradually aggregate as they disperse into an aquatic environment.¹⁰ The dispersed morphology of the MWCNTs supported the fact that they did not totally stretch out as polyline but got tangled like a loose ball of string with diffusing ends as they dispersed into the wastewater (Fig. S3†). The MWCNTs adsorbed and intertwined on the sludge particles as they were dispersed with the activated sludge in the SBR (Fig. S4†). However, it should be pointed out that the morphology of the MWCNTs did not change significantly with the increasing dispersal time. This indicated that steric repulsion impacted the adsorption of the organic matter in the wastewater and kept the initial morphology of the MWCNTs. This was in agreement with other studies.⁴³ Besides, the MWCNTs showed some adsorbed ability onto antibiotics in aquatic environment. The concentration of AMP in the water (10 mg L⁻¹) decreased to 4.627 mg L⁻¹ as the MWCNTs (1 g L⁻¹) were applied as adsorbents (Fig. S5†).

3.2 Operating state of the SBR

The number of microorganisms, respiration intensity, and effluent quality are important indicators for the operating state of the sewage treatment; therefore, the MLVSS, SOUR, effluent TOC and AMP concentration of various SBR were determined and are shown in Fig. 2.

Fig. 2 Co-effect of AMP and MWCNTs on (A) MLVSS; (B) SOUR; (C) effluent TOC value; and (D) effluent AMP concentration of SBR. Error bars represent standard deviations determined from measurements performed in triplicate.

It can be found in Fig. 2A that the MLVSS values in R_CNT, R_AMP and R_D were lower than 100% of R_CP for the 24 h exposure case. This may be due to the cytotoxicity of AMP and MWCNTs. AMP can induce cell shrinkage, apoptosis and necrocytosis through inhibiting cell wall synthesis and cell granulation, and MWCNTs exhibit a strong antimicrobial activity via oxidative stress and membrane perturbation.^5,44 However, the MLVSS values of various R_D were larger than those of R_CNT and R_AMP, and increased with the increase in the concentration of MWCNTs. This suggests that compared with exposure to individual AMP or MWCNTs, higher amounts of original microorganisms could resist the co-occurrences of AMP and MWCNTs. A similar phenomenon was found in other reports; whereby an antagonistic effect could be observed between the toxicity of carbon nanoparticles and toxic pollutant (e.g. phenol and Cu²⁺).⁴⁵ This was mainly due to the occupation of the functional groups of the AMP and MWCNTs in the adsorption. MWCNTs could bind AMP in wastewater through –OH, –COOH and so on. The adsorbed AMP would twist its functional groups and then could hardly bind to penicillin binding proteins,⁴⁶ which is a key step to inhibiting the bacterial cell wall synthesis of AMP.⁴⁷ Thus, the effect of AMP on bacteria would be weakened. Meanwhile, the adsorption reaction between AMP and MWCNTs would result in the occupation of the functional groups of MWCNTs, which has been proved to be related with the biotoxicity of CNTs. Gao et al. found that the surface chemistry modification on MWCNTs can cause less immune perturbations in mice and in macrophages through altering the receptor recognition.⁴⁸ Thus, the effect of MWCNTs on bacteria also would be restrained. After 240 h exposure, the surviving microorganisms adapted to the co-occurrences, whether MWCNTs, AMP or AMP combined with MWCNTs, and this was supported by the MLVSS values of the various SBRs, which were nearly recovered to the level of R_CP (98.6–101.6%).

Besides the number of active microorganisms, microbial respiration was influenced by AMP and MWCNTs based on the SOUR analysis (Fig. 2B). The SOUR of R_CNT was slightly improved to 112% of R_CP, while the SOUR of R_AMP evidently declined to 78% of SR_CP after 24 h exposure. This indicated that the MWCNTs could enhance microbial respiration, although it resulted in the death of some microorganisms. In agreement with other reports, it was found that CNTs could increase the activities of some microorganisms and enzymes through promotion of the generation, transfer and consumption of electrons.⁴⁹ The co-effect of AMP and MWCNTs on the respiration intensity of microorganisms was similar to that of individual MWCNTs for 24 h exposure. Moreover, the SOUR of R_D increased with the increase in the concentration of MWCNTs. The microbial respiration of R_D was even slightly higher than that of R_CNT when the concentration of MWCNTs was over 50 mg L⁻¹. Although the surviving microorganisms adapted to the presence of AMP, MWCNTs, or AMP combined with MWCNTs in the 240 h exposure case, the SOUR values of R_CNT, R_AMP or R_D decreased to 50–60% of R_CP.

The TOC values of the effluent fluctuated from between 3.3 and 3.7 mg L⁻¹ during the whole exposure experiment, in which the AMP concentrations showed a decreasing trend and then an upward trend with the operation time, especially for R_D (Fig. 2CD and ). Compared to Fig. 2B, it could be found that the AMP efficiency of wastewater treatment in SBR is closely related to the microbial respiration. The effluent quality would gradually deteriorate with SOUR decreasing. Also, it should be noted that in the first five days, the AMP concentrations in the effluent of R_D were lower than that of R_AMP. This may be due to the adsorption of MWCNTs and the high metabolic viability of the bacteria. The MWCNTs showed a high adsorption ability for AMP (Fig. S5†), which would lead to a decrease in the concentration of AMP in R_D with the increase in the concentrations of MWCNTs within the first five days. Moreover, the boosted metabolic viability of the aerobic heterotrophic bacteria in R_D (the highest SOUR) also contributed to the lower AMP concentration in the effluent. Besides, the highest content of polysaccharide, whose adsorption is the primary path for the removal of antibiotics in activated sludge process,^50–52 in R_D was one of the contributors to the low concentration of AMP in the effluent (Fig. S6†). However, the quality of the effluent for R_D sharply deteriorated after seven days of exposure due to the decrease in SOUR and the adsorption saturation of the MWCNTs, especially for concentrations of MWCNTs of 20 mg L⁻¹ and 50 mg L⁻¹.

3.3 Physiological status of the microorganisms

The operational efficiencies of SBR are related to the physiological status of the microorganisms, including intact, necrotic and lysed cells, which is directly influenced by the cytotoxicity of the AMP and MWCNTs.^5,34,53,54 In this study, the physiological status of the microorganisms in the SBR was evaluated according to the scheme shown in Fig. 1, and the results are shown in Fig. 3–5.

Fig. 3 Different membrane permeabilizations of microbial cells in R_CP, R_CNT, R_AMP and R_D: (A) after 24 h exposure and (B) after 240 h exposure. (a) R_CP; (b) R_CNT; (c) R_AMP; (d) R_D20; (e) R_D50; (f) R_D100, and quadrant P (PI positive) represents the cell membrane permeabilities.

Fig. 4 Relative fold changes in cellular metabolic and proliferative viability in R_CP, R_CNT, R_AMP and R_D. Error bars represent the standard deviations determined from measurements performed in triplicate.

Fig. 5 Microbial cell apoptosis and/or necrosis in R_CP, R_CNT, R_AMP and R_D: (A) after 24 h exposure and (B) after 240 h exposure. (a) R_CP; (b) R_CNT; (c) R_AMP; (d) R_D20; (e) R_D50; (f) R_D100.

Membrane permeability, which can be determined through PI staining, is associated with the integrity of the cell membrane.^55–57 FCM analysis showed that the membrane permeability was increased in R_AMP and R_CNT (as evidenced by the 3.92- and 3.43-fold increase after 24 h exposure, respectively). This agreed with other studies that found that CNTs could damage the cell membrane by oxidative stress and membrane agitation, while AMP could result in cell debris due to DNA degradation or cell granulation.⁵ In contrast, the membrane permeability in R_D was significantly lower than that in R_AMP and R_CNT and decreased with the increasing concentration of MWCNTs. This suggests that the presence of enough MWCNTs would protect microbes from the inflammatory effect of AMP. Because more microbes could maintain their integrated cell membrane to AMP combined with MWCNTs, the MLVSS values of R_D were larger than those of R_CNT and R_AMP for the 24 h exposure case.

However, cell membrane integrity is thought not to fully represent cell activity because some of the cellular functions would cease before the membrane integrity is lost.⁵⁸ Thus, the cell metabolic and proliferative viability were analysed based on the fluorescence-based readout of the Alamar Blue (AB) assay, which is highly sensitive to the proliferation of microorganisms.⁵⁹ As shown in Fig. 4, it could be found that the cell metabolic viability in R_AMP, R_CNT and R_D decreased slightly compared with that of R_CP after 24 h exposure based on the fluorescent signal produced by microbial proliferation, except when the concentration of MWCNTs was 100 mg L⁻¹. For 240 h exposure, the metabolic viability in R_AMP, R_CNT and R_D was further reduced. Compared with R_AMP, R_D showed a weaker inhibition of the metabolic viability, irrespective of whether it was 24 h or 240 h exposure. Furthermore, the metabolic viability of the activated sludge increased relatively with the increase in the concentration of the MWCNTs. The metabolic viability in R_D100 was even higher than that of R_CP after 24 h exposure. This higher metabolic viability of the activated sludge contributed to a higher SOUR of R_D for the 24 h exposure case (Fig. 2B).

Generally, cell metabolic and proliferative viability are closely related to cellular reactive oxygen species (ROS) generation, which are natural by-products during the process of normal aerobic metabolism, especially when the cells interact with exogenous pollutants.⁶⁰ Excessive ROS generation could destroy the dynamic equilibrium between the oxidant and anti-oxidant processes, and damage DNA, protein, and lipids in microbial cells.⁵³ This causes a cumulative inhibition effect on the microbial respiratory activity, which then leads to irreparable metabolic dysfunction.^53,61R_AMP showed the highest ROS production (3–6-fold that of R_CP) in this study (Fig. S7†). However, the cellular ROS level (1.5–2-fold that of R_CP) was significantly lower than that of R_AMP, and similar to that of R_CNT (1.5–1.8-fold that of R_CP) as AMP and MWCNTs co-existed in the SBR. Thus, this may be one of the reasons why the proportion of cell metabolic and proliferative viability in R_D was higher than that in R_AMP.

Beside the ROS, the cell metabolic and proliferative viability were found to be associated with extracellular polysaccharide as well. As a complex high-molecular-weight mixture of polymers produced by microorganisms,⁶² extracellular polysaccharide can play an essential role in the removal of trace contaminants through absorbing many organic pollutants, including antibiotics, through hydrophobic regions.^63,64 Furthermore, extracellular polysaccharide could provide a formidable protective defence against any interference from the external environment.^65,66 Previous studies have proved that NPs could be prevented by extracellular polysaccharide from piercing through the membrane; thereby averting cytotoxicity to microorganisms.⁶⁷ As AMP and MWCNTs co-existed in the SBR in our experiment, the largest amount of extracellular polysaccharide would be observed in R_D (Fig. S6†).

This would reduce the cytotoxicity of MWCNTs to the microorganisms, which may be another reason for the higher cell metabolic and proliferative viability in the SBR in the co-presence of AMP and MWCNTs.

Excessive ROS generation ultimately contributed to a higher number of dead bacteria in R_CNT, R_AMP and R_D compared to R_CP (Fig. S8†). However, the production of the largest amount of extracellular polysaccharide contributed to the higher number of viable bacteria with an intact membrane and a lower number of dead bacteria in R_D compared to R_AMP and R_CNT (Fig. S8†). Combining the effect of the ROS and extracellular polysaccharide, the co-occurrence of AMP and MWCNTs showed a lower cytotoxicity compared with the individual AMP or MWCNTs. This contributed to higher MLVSS values of R_D than R_AMP and R_CNT for the 24 h exposure (Fig. 2).

In order to further investigate the co-effect of AMP and MWCNTs, the cell death modes were adopted. The cell death modes within activated sludge were divided into: early apoptotic cells, late apoptotic cells, and necrotic cells attributed to cell shrinkage, phosphatidylserine (PS) exposure and DNA fragmentation.^32,33,68 These were quantified based on AV-FITC/PI staining, and the results are presented in Fig. 5, which shows the live cells (AV⁻ PI⁻), early apoptotic cells (AV⁺ PI⁻), late apoptotic cells (AV⁺ PI⁺) and necrotic cells (AV⁻ PI⁺) distributions.

For the 24 h exposure, the proportions of late apoptotic and necrotic cells increased dramatically in R_AMP, and that of the early apoptotic cells increased in R_CNT compared to R_CP. The higher proportions of early apoptotic cells, late apoptotic cells and necrotic cells in R_CNT and R_AMP contributed to a decrease in the MLVSS values for the 24 h exposure. Furthermore, the proportion of necrotic cells in R_D was significantly higher than that in R_CP, indicating that the co-occurrences of AMP and MWCNTs would mainly produce cytotoxicity and genotoxicity based on DNA breakage and fragmentation. However, the proportion of necrotic cells in R_D decreased gradually with the increasing concentration of MWCNTs for the 24 h exposure (Fig. 5A). For example, when the concentration of MWCNTs increased to 100 mg L⁻¹, the proportion of necrotic cells in R_D was similar to that in R_CP.

The results for the physiological status of the microorganisms showed that the co-occurrence of AMP and MWCNTs had lower cytotoxicity compared with the individual AMP or MWCNTs after 24 h exposure. For 240 h exposure, the proportion of necrotic cells in R_D increased dramatically (from 2.00- to 3.29-fold), and the proportions of both apoptotic and necrotic cells increased sharply in R_CNT compared to those in R_CP (Fig. 5B). This was also supported by the higher number of dead bacteria in R_CNT, R_AMP and R_D with the increase in exposure time (Fig. S8†). The large amount of apoptotic and necrotic cells in R_CNT, R_AMP and R_D inhibited the cell metabolic viability in the activated sludge, which led to a sharp decrease in SOUR for the 240 h exposure.

3.4 Microbial community structure and ARGs analysis

The system stability and wastewater treatment performance of activated sludge exposed to antibiotics depend mostly on the microbial community structure.⁶⁹ The microbial community structure in R_AMP, R_CNT and R_D was visualized by UPGMA clustering tree, histogram and a heatmap of the classified genera (Fig. 6A–C). The UPGMA clustering tree showed that the microbial community structure in R_AMP, R_CNT and R_D deviated from the R_CP due to the effect of the AMP and MWCNTs (Fig. 6A). The community structure in R_CNT was obviously different from that of R_AMP and R_D. R_AMP showed almost the same community structure as R_D when the concentration of MWCNTs was 20 mg L⁻¹. However, the community structures of R_D gradually deviated from R_AMP as the concentration of MWCNTs increased to 100 mg L⁻¹.

Fig. 6 Microbial community structure and relative abundance of ARGs in R_CP, R_CNT, R_AMP and R_D. (A) UPGMA clustering tree based on the unweighted unifrac distance (distribution of phyla); (B) analysis of taxonomic composition; (C) heatmap of the classified genera (top 20) and (D) the relative abundance of ARGs. Error bars represent the standard errors of the duplicate biological samples.

Analysis of the taxonomic composition at the phyla level, as shown in Fig. 6B, indicated the presence of Proteobacteria, Acidobacteria, Bacteroidetes, Chloroflexi, Planctomycetes, Actinobacteria and Gemmatimonadetes, which are commonly found in wastewater treatment systems,⁷⁰ and these were the dominant phyla in R_CP. Acidobacteria, Chloroflexi, Actinobacteria and Nitrospirae tended to increase, whereas Proteobacteria, Bacteroidetes, Planctomycetes and Gemmatimonadetes decreased, while Patescibacteria phylum emerged after 240 h exposure to MWCNTs. The exposure to AMP did not result in a decrease in Proteobacteria. The taxonomic composition in R_D was similar to that in R_AMP when the concentration of MWCNTs was 20 mg L⁻¹, and it tended to be similar to that of R_CNT with increasing concentrations of MWCNTs. However, it is worth noting that Firmicutes phylum emerged in the top 20 phyla in R_D50 and R_D100, but not in the other exposures.

To further illustrate the response of the microbial compositions in the co-occurrences of AMP and MWCNTs, the populations at the genus level are identified in Fig. 6C. Exposure to individual MWCNTs greatly increased the abundance of Dechloromonas, Zoogloea, Legionella and Bdellovibrio, while AMP led to a significant increase in the abundance of Defluviimonas and Chlorobi bacterium OLB5. The genera with great increases in abundance after exposure to the co-occurrences of AMP and MWCNTs were Bacteroide, Alloprevotella, Thauera, etc., especially in R_D50 and R_D100. The different microbial community structures induced by the exposure of AMP and MWCNTs would be another reason for the cell metabolic and proliferative viability in the SBR.

It has been reported that the host bacteria contribute to the dominant gene type, and also, the resistant microorganism fates.^71,72 Herein, the relative abundance (defined as the absolute copy number of genes normalized to the absolute copy number of ambient 16S-rRNA genes) was used in this study to compare the differences in three AmpC β-lactamase genes (bla_FOX, bla_DHA and Lak) and two integron (intI1 and intI2) acquired in R_CP, R_AMP and R_D after 240 h exposure, and the results are shown in Fig. 6D. Both R_AMP and R_D showed enrichment in all detected genes compared to R_CP due to the presence of antibiotics. Also, the relative abundances of bla_DHA, Lak, intI1 and intI2 in R_D were significantly (p < 0.05) higher than those detected in R_AMP. This indicated that the co-occurrences of AMP and MWCNTs would further increase the abundance of ARGs in the activated sludge. In addition, a high relative abundance of intI1 and intI2, which are considered as potential indicators for ARGs horizontal transfer,⁷³ in R_D meant that the MWCNTs showed a synergistic effect on ARGs horizontal transfer when they were combined with antibiotics.

Non-susceptible bacteria may survive exposure to antibiotics, with a consequence on the abundance of resistance genes during wastewater treatment.⁷⁴ In the meanwhile, bacteria without ARGs could convert into ARB through ARGs horizontal transfer.⁷⁵ The ARGs horizontal transfer is thought to be the main reason for the increase in the abundance of ARGs, and this could co-occur with the removal and accumulation of antibiotics during the wastewater treatment process.⁷⁶ The process of ARGs transfer, replication and formation could not continue unless the antibiotics concentrations in the cell could be reduced below the threshold concentration.⁷⁷ The concentrations of AMP in R_D were significantly lower than those in R_AMP for 24 h exposure (Fig. 2D). Consequently, there were more opportunities for those bacteria without ARGs to survive and acquire resistance genes in the initial exposure period. In addition, MWCNTs would still increase ARGs conjugative transfer frequency as well as facilitate the microbial uptake of naked resistance plasmids based on a modest rise in membrane permeability, like nano-ZnO and nano-Al₂O₃ in previous reports.^19–21

It should be pointed out that our research was specific to AMP and MWCNTs, and whether the relative mechanism on antibiotics with other NPs are acceptable is unknown. In addition, considering the toxicity of NPs in natural and engineered aquatic systems is inseparable from considering the environmental variables, including natural organic matter, ionic strength and aggregation state. These are all important to deeper research on antibiotics and NPs to explore their potential mechanisms in the future.

4. Conclusions

This study investigated the microbial physiological status, microbial community structure and resistance genes in SBRs in order to elaborate the co-effects of AMP and MWCNTs on activated sludge in WWTPs. The co-occurrences of AMP and MWCNTs showed a lower cytotoxicity to activated sludge in SBRs compared to individual ones. The microorganisms in R_D showed a lower membrane permeability and higher cell metabolic viability than that in R_AMP and R_CNT. This resulted in lower AMP concentrations and TOC values of the effluent of R_D in the initial exposure period with a higher MLVSS and SOUR values compared with R_AMP and R_CNT. After 240 h exposure, the quality of effluent for R_D sharply deteriorated, although the surviving microorganisms adapted to the exposure to AMP and MWCNTs. Also, it should be noted that the relative abundance of ARGs in R_D was significantly higher than that in R_AMP after 240 h exposure. These conclusions indicated that the long-term exposure to AMP and MWCNTs would magnify the dissemination risk of ARB and ARGs.

Conflicts of interest

There are no conflicts to declare.

Acknowledgements

This study was financially supported by Shandong Provincial Natural Science Foundation (no. ZR2019MEE006) and the grants from Tai Shan Scholar Foundation (no. ts 201511003). We would like to thank Zhifeng Li, Jingyao Qu from State Key Laboratory of Microbial Technology of Shandong University for assistance in ImageStreamX Mark II flow cytometry (Amnis, USA). We also thank TopEdit (www.topeditsci.com) for its linguistic assistance during the preparation of this manuscript.

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Footnote

† Electronic supplementary information (ESI) available. See DOI: 10.1039/d0en00668h

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