Chemosphere 135 (2015) 138–145

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Fate of antibiotic resistant cultivable heterotrophic bacteria and antibiotic resistance genes in wastewater treatment processes Songhe Zhang a,b,⇑, Bing Han a, Ju Gu a, Chao Wang a, Peifang Wang a,⇑, Yanyan Ma a, Jiashun Cao a, Zhenli He b a b

Ministry of Education Key Laboratory of Integrated Regulation and Resource Development on Shallow Lakes, College of Environment, Hohai University, Nanjing 210098, China University of Florida, Institute of Food and Agricultural Sciences, Indian River Research and Education Center, 2199 South Rock Road, Fort Pierce, FL 34945, USA

h i g h l i g h t s

g r a p h i c a l a b s t r a c t

 WWTPs contributed to removal of

cultivable heterotrophic bacteria from sewage.  Gram-negative and -positive bacteria dominated in the influent and effluent, respectively.  The mean MAR index was lower in the influent than effluent samples.  ARG abundance increased in the activated sludge after WWTP processes.

a r t i c l e

i n f o

Article history: Received 22 September 2014 Received in revised form 1 April 2015 Accepted 2 April 2015

Handling Editor: Shane Snyder Keywords: Gram-negative Influent Effluent Multi-antibiotic resistance index Genera

a b s t r a c t Antibiotic resistant bacteria (ARB) and antibiotic resistance genes (ARGs) are emerging contaminants of environmental concern. Heterotrophic bacteria in activated sludge have an important role in wastewater treatment plants (WWTPs). However, the fate of cultivable heterotrophic ARB and ARGs in WWPTs process remains unclear. In the present study, we investigated the antibiotic-resistant phenotypes of cultivable heterotrophic bacteria from influent and effluent water of three WWTPs and analysed thirteen ARGs in ARB and in activated sludge from anoxic, anaerobic and aerobic compartments. From each influent or effluent sample of the three plants, 200 isolates were randomly tested for susceptibility to 12 antibiotics. In these samples, between 5% and 64% isolates showed resistance to >9 antibiotics and the proportion of >9-drug-resistant bacteria was lower in isolates from effluent than from influent. Eighteen genera were identified in 188 isolates from influent (n = 94) and effluent (n = 94) of one WWTP. Six genera (Aeromonas, Bacillus, Lysinibacillus, Microbacterium, Providencia, and Staphylococcus) were detected in both influent and effluent samples. Gram-negative and -positive isolates dominated in influent and effluent, respectively. The 13 tetracycline-, sulphonamide-, streptomycin- and b-lactam-resistance genes were detected at a higher frequency in ARB from influent than from effluent, except for sulA and CTX-M, while in general, the abundances of ARGs in activated sludge from two of the three plants were higher in aerobic compartments than in anoxic ones, indicating abundant ARGs exit in the excess sledges and/or in uncultivable bacteria. These findings may be useful for elucidating the effect of WWTP on ARB and ARGs. Ó 2015 Elsevier Ltd. All rights reserved.

⇑ Corresponding authors at: Ministry of Education Key Laboratory of Integrated Regulation and Resource Development on Shallow Lakes, College of Environment, Hohai University, Nanjing 210098, China (S. Zhang). E-mail addresses: [email protected] (S. Zhang), [email protected] (P. Wang). http://dx.doi.org/10.1016/j.chemosphere.2015.04.001 0045-6535/Ó 2015 Elsevier Ltd. All rights reserved.

S. Zhang et al. / Chemosphere 135 (2015) 138–145

1. Introduction Numerous antibiotics have been extensively used to treat infectious diseases in humans and to promote the growth of food animals in agriculture (da Costa et al., 2008; Martinez, 2008). Residual antibiotics and their metabolites can be released from the body into the environment, causing chemical pollution (Gao et al., 2012). The practices of antibiotic use have influenced all the aspects of microbial genetic ecology, as revealed by the fact that antibiotic-resistant bacteria (ARB) and antibiotic resistance genes (ARGs) have been frequently detected in both liquids (wastewater, surface water, groundwater and even drinking water) (Schwartz et al., 2006; Storteboom et al., 2010) and in solids (sludge, soil and sediment) from the environment (Auerbach et al., 2007; Yang et al., 2014). Selective pressure from antibiotics can not only increase the concentrations of ARB by inhibiting antibiotic-susceptible bacteria, but also contributes to the selection of mutations and to horizontal gene transfer (HGT) between microorganisms (Martinez, 2008; Wang and Schaffner, 2011). Given that ARB and ARGs of clinical concern have been documented to arise from environmental sources (Martinez, 2008), greater attention has been focused on limiting resistance dissemination pathways between humans and the environment (Baquero et al., 2008). Wastewater treatment plants (WWTPs) are supposed to hold an important place in the reduction or spreading of antibiotics and ARB (Iwane et al., 2001; Guardabassi et al., 2002; Munir et al., 2011). Although neither the links between the presence of residual antibiotics in WWTPs and the favouring of resistant bacteria nor the transfer of resistance at the low antibiotic concentrations in the environment have been well established (Martinez, 2008; Bouki et al., 2013), accumulated data show that WWTPs potentially provide an environment that has potential for the development and/or spread of resistance, as bacteria are continuously mixed with antibiotics at sub-inhibitory concentrations (Aminov et al., 2001; Auerbach et al., 2007; Davies, 2012). In addition, other factors in WWTPs, such as gene cassettes, integrons, plasmids, and heavy metals, play important roles in the exchange of resistance and contribute to resistance retention and dissemination in the WWTP systems (Stepanauskas et al., 2006; Zhang et al., 2011; Moura et al., 2012). In WWTPs, activated sludge containing relatively stable, highly complex microbial communities has an important role in the removal of nutrients and pathogens from sewage. A recent report showed that there is a difference in the bacterial community structure and composition between aerobic and anaerobic sludge (Wang et al., 2013b). Activated sludge usually contains uncultivable bacteria that are difficult to study by traditional molecular methods, and 16S rRNA gene sequences from many bacteria cannot even be assigned to taxonomic groups at the phylum or class levels (Zhang et al., 2012). Currently, the majority of studies and reviews focusing on antibiotic resistance in WWTPs give special emphasis to faecal indicators and pathogens (including faecal coliforms (e.g. Enterobacter, Klebsiella and Citrobacter), enterococcal species, Acinetobacter spp. and Staphylococcus spp.) that are usually used to assess microbiological water quality and safety (Chen and Zhang, 2013; Łuczkiewicz et al., 2010; Korzeniewska et al., 2013; Rijal et al., 2009). Current studies of antibiotic resistance in bacteria have typically focused on pathogens and faecal indicators such as Enterobacteriaceae, Aeromonas spp. (Igbinosa and Okoh, 2012), Escherichia coli (Reinthaler et al., 2003) and Acinetobacter (Zhang et al., 2009), which only represent a small proportion of the total cultivable heterotrophic bacteria. Though pathogens do not produce antibiotics, they can harvest ARGs released from other

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bacteria in environment. However, minimal is known about the characteristics and species of antibiotic-resistant heterotrophic bacteria, which dominate in activated sludge (Reinthaler et al., 2003; Ferreira da Silva et al., 2007). Since most of the heterotrophic bacteria are uncultivable (Yang et al., 2014), qPCR methods are generally employed to investigate the ARGs at the samples level, including in samples from anaerobic and aerobic environment (Ma et al., 2011; Wang et al., 2013b) and WWTP sludge (Yang et al., 2014). However, the alterations in ARG concentrations under different bio-treatment processes remain unclear. Since the bacteria can obtain ARGs from the environment through horizontal gene transfer (HGT) and the sewage and sludge can be used on farm (Joao and Eddie, 2013), it is necessary to investigate effects of WWPTs on the ARGs and ARB. The aims of the present study were to: (1) evaluate the alterations in antibiotic resistance of cultivable heterotrophic bacteria; (2) study whether the ARGs in the cultivable ARB from influent were different from those in effluent; and (3) investigate whether ARGs increased in compartments with the water flow. To accomplish these aims, the concentrations of cultivable heterotrophic bacteria were determined in the influent and effluent of three full-scale WWTPs (WWTPa, WWTPb, and WWTPc), and the susceptibility of these heterotrophic bacteria to 12 antibiotics was evaluated. The frequencies and concentrations of 13 ARGs were determined in the randomly selected isolates from influent and effluent, and in activated sludge samples from anaerobic, anoxic and aerobic compartments. 2. Materials and methods 2.1. Characteristics of wastewater treatment plants Samples of influent, effluent and activated sludge were collected from WWTPa, WWTPb and WWTPc) located in Wuxi, Jiangsu province, China. These plants have anaerobic, anoxic and aerobic treatment processes (Fig. S1). WWTPa has a secondary sedimentation tank, while a membrane bioreactor (MBR) is employed to replace the secondary sedimentation tank in WWTPb and WWTPc. WWTPa and WWTPb receive wastewater containing at least 75% domestic sewage and 25% pretreated industrial effluent, while WWTPc receives 70% domestic sewage and 30% pretreated industrial effluent. The parameters of these plants were provided in Table S1. 2.2. Sample collection The influent samples (2 L) were collected after passing through the grille (Fig. S1). The effluent samples (2 L) were collected from the influent from sedimentation or MBR but not exposed to ultraviolet radiation or oxidizer (Fig. S1). Additionally, 3 L of activated sludge samples were collected from the anaerobic, anoxic and aerobic tanks of WWTPa and WWTPb. All these samples were collected in March (dry season) or July (wet season) 2012. Samples were kept on ice during transportation to the laboratory. Water samples of influent and effluent were immediately processed within 12 h. Sludge samples were centrifuged at 8000 rpm for 10 min at 4 °C, and the pellets were stored at 80 °C until DNA extraction. 2.3. Enumeration of cultivable heterotrophic bacteria and total coliforms Total heterotrophs and total coliforms in the water samples were quantified by a dilution plating procedure (Korzeniewska et al., 2013). The samples were serially diluted using phosphate-

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buffered saline (PBS, pH = 7.2 ± 0.2), and 0.1 ml of the serial dilution was used for plate inoculation. Cells were plated on R2A agar and MacConkey nutrient agar to count the numbers of total heterotrophic isolates and total coliforms, respectively. R2A plates were incubated at 28 °C for 36 h, while MacConkey plates were incubated at 37 °C for 24 h. For each assay, triplicate plate counts were performed with five different dilutions for each sample. 2.4. Antibiotic susceptibility test For each sample, 200 isolates (about 1–5% of total isolates) were randomly selected from the R2A agar plates and were further purified by the plate streak method. According to CLSI, 2005 guidelines, the purified isolates were cultured on Mueller–Hinton agar (Oxoid) containing none (control) or one of 12 antibiotics listed below: ampicillin (AMP, 32 lg mL 1), cefazolin (CFZ, 32 lg mL 1), ceftriaxone (CFR, 64 lg mL 1), tetracycline (TET, 16 lg mL 1), oxytetracycline (OTC, 40 lg mL 1), nalidixic acid (NA, 32 lg mL 1), levofloxacin (LEV, 8 lg mL 1), gentamicin (GEN, 16 lg mL 1), streptomycin (STR, 32 lg mL 1), sulfamethoxazole (SMZ, 76 lg mL 1), vancomycin (VAN, 32 lg mL 1) or chloramphenicol (CHL, 32 lg mL 1). The concentrations of these antibiotics used in this study were according to the recommended concentrations in CLSI, 2005 guidelines (For details see supplementary materials). Two antibiotic-sensitive organisms, Staphylococcus aureus ATCC 29213 and E. coli ATCC 25922, were used as representatives of Gram-positive and Gram-negative bacteria, respectively. After incubation in the dark at 28 °C for 2 days, clear colony formation was recorded as resistance if no S. aureus ATCC 29213 or E. coli ATCC 25922 colonies were observed on the same plate. If an isolate did not grow well on the Mueller–Hinton agar without any antibiotic, R2A agar plates were used for antibiotic screening. All the experiments were performed at least twice. The antibiotic resistance frequency was calculated using the equation (a/b) ⁄ 100%, where a is the number of isolates resistant to antibiotics and b is the number of isolates from the sample. The multiple antibiotic resistance (MAR) index was calculated to evaluate the potential resistance of isolates to multiple antibiotics (Wang et al., 2013a). The MAR index was calculated as a/(b  c), where a is the aggregate resistance score of all isolates from one sample, b is the number of antibiotics tested, and c is the number of isolates. 2.5. Detection of ARGs and genus (species) identification of ARB To identify the genera or species of these heterotrophic isolates and to assay their antibiotic-resistance genotypes, isolates from WWTPb in July were further analysed. A total of 188 of the aforementioned purified bacterial isolates were randomly selected from influent (n = 94) and effluent (n = 94). These isolates were cultured in 5 ml liquid Luria–Bertani media in shaking flask at 37 °C for 16 h. A 1.5-ml volume of each cultured isolate was used for DNA extraction using the methods described in a previous study (Wang et al., 2013a). The universal bacterial primers 8f (AGAGTTTGATCCTGGCTCAG) and 1492r (GGTTACCTT GTTACGACTT) were used for PCR amplification (Krause et al., 1999). The PCR was performed with initial denaturation at 94 °C for 5 min, 30 cycles at 94 °C for 60 s, 53 °C for 60 s and 72 °C for 30 s with a final extension step at 72 °C for 10 min. The PCR products were cloned and transported to a company (Sangon Biotech Co. Ltd., Shanghai, China) for DNA sequencing. The resulting 16S rRNA gene sequences were queried against the GenBank database using the BLAST alignment tool (http://www.ncbi.nlm.nih.gov/ blast/). The species were identified according to known species name with the highest score value.

The extracted DNA was stored at 20 °C for subsequent analysis. The presence of 13 antimicrobial resistance genes was examined by PCR with gene-specific primers. The primers for the ARGs were provided in Table S2. PCR amplifications were conducted in a volume of 25 lL containing 1 PCR buffer, 1.5 mM MgCl2, 100 lM dNTP, 3 pmol of each primer, 1 lL of template DNA, 1 U of Taq polymerase and distilled water. Amplification conditions were as follows: initial denaturation at 94 °C for 5 min, 30 cycles at 94 °C for 60 s, annealing for 60 s at the annealing temperature (Table S2), extension at 72 °C for 30 s, and a final extension step at 72 °C for 10 min. Triplicate PCR reactions were performed for each sample, and double-distilled water was used as a negative control with each primer set. PCR was performed in a Bio-Rad S1000 thermal cycler (Bio-Rad, USA). The products were analysed by electrophoresis using 0.81.0% agarose gel with golden view in 1 TAE buffer at 120 V for 20 min and visualisation under UV transillumination. 2.6. DNA extraction of sludge and real-time qPCR of ARGs Total DNA was extracted from 0.30 g sludge using PowerSoil kits (MO BIO Laboratories, Inc. USA) in accordance with the manufacturer’s protocol. Real-time qPCR was employed to quantify the levels of ARGs. DNA extracts were quantified using a NanoDrop spectrometer (ND-1000, Thermo Fischer Scientific, Wilmington DE, USA) and a serial dilution (1:50–500) of DNA extracts were diluted to an optimal concentration for real-time qPCR. Reactions were assembled in 48-well plates with a final volume of 20 lL using SYBR Green Super Mix (BioRad, Hercules, CA, USA). The primers for the ARGs and bacterial 16S rRNA genes are listed in Table S2. The qPCR reactions were performed with a program of 5 min at 95 °C, 50 cycles of 15 s at 95 °C; 30 s at the annealing temperature (Table S2) incorporated with a melting curve stage with temperature ramping from 55 °C to 95 °C and an extension step at 72 °C for 30 s. The DNA of these target genes from ARBs were cloned into the pGEMÒ-T Easy plasmids (Promega Promega Corporation, Madison, WI, USA) and 10-fold serial dilutions of plasmids ranging from 108 to 102 gene copies lL 1 were used to construct qPCR standard curves of these genes. The R2 values were more than 0.99 for all calibration curves. Samples were analysed in triplicate with a standard curve, and a negative control was included in each run. Relative concentrations of ARGs (normalised to the 16S rRNA gene copy number) were calculated in each activated sludge sample. Gene concentrations were quantified in triplicate within each assay. The partial 188 16s rRNA sequence had been submitted to the Genbank at NCBI site (NO. —). 2.7. Statistical analysis The concentrations of total heterotrophic isolates and faecal coliforms were log10-transformed and were expressed as the mean ± S.D. after log10 transformation. The concentrations of ARGs in activated sludge were expressed as the mean ± S.D. The concentrations of ARGs in anoxic, anaerobic and aerobic compartments from the same plant at the same sample time were compared using one-way analysis of variance (ANOVA) followed by Tukey’s post hoc analysis at P < 0.05. 3. Results 3.1. Concentrations of total heterotrophs and total coliforms in influent and effluent samples The log10 values of total heterotroph concentrations ranged from 5.69 to 6.53 in the influent and from 3.01 to 5.26 in the

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effluent samples, whereas total coliforms concentrations ranged from 4.07 to 4.68 in the influent and from 1.60 to 3.18 in the effluent (Table 1). The removal rates of total coliforms ranged from 98.4% to 99.9% in July and ranged from 96.8% to 99.3% in March, 2012. 3.2. Antibiotic susceptibility of heterotrophs isolated from effluent and influent samples To evaluate the resistance of these cultivable heterotrophic bacteria, 200 isolates were isolated from each influent or effluent water sample of the three plants to identify their susceptibility to 12 antibiotics. The resistance of AMP, CFZ, NA, SMZ, CHL, OTC, GEN, CFR, TET and STR occurred in the isolates in the proportions ranging from 53% to 96%, while CFR and LVE resistance occurred at 40% and 13%, respectively (Table 2). The mean MAR indices of effluent samples of WWTPc ranged from 0.53 (March) to 0.78 (July) (Table 2). Heterotrophic isolates showing resistance to more than 5 antibiotics accounted for 60% (influent of WWTPb in July) to 98% (effluent of WWTPa and WWTPc in March) (Fig. 1). The percentage of isolates showing resistance to at least 10 drugs ranged from 5% (effluent of WWTPa in July) to 64% (effluent of WWTPc in March) and was higher in influent than effluent for the same plants at the same sampling time.

28% and 19% of the influent isolates, respectively, as compared with 26%, 26%, 10% and 20% of the effluent isolates (Table 4). Four TET resistance genes, tetA, tetB, tetE and tetO were detected in the TET-resistant bacteria, and the frequencies of tetO and tetE detection were higher than those of tetA and tetB in both the influent and effluent samples. The presence of the four tet genes explained 92% and 78% of the tetracyclines antibiotic resistance in the influent and effluent samples, respectively. In streptomycin-resistant bacteria, the frequencies of str genes (Table 4) were higher in Gram-negative genera than Gram-positive genera, and strA and strB occurred in the influent at 55% and 35%, respectively, and in the effluent at 25% and 18%, respectively. Four b-lactamase genes were detected in all the genera except for Pannonibacter Microbacterium and Staphylococcus. TEM (>67%) was the most dominant extended-spectrum beta-lactamase (ESBL) gene in b-lactam-resistant isolates, and it occurred in 80% of Aeromonas isolates. For a given ARG, the harbouring frequencies were higher in Gram-negative ARB as compared to Gram-positive ARB in both the influent and effluent samples, with the exception of tetO (Table 3). The mean MAR indices of the Gram-negative isolates were higher than those of Gram-positive isolates in the same samples, and the mean MAR indices of the Gram-negative or -positive isolates were lower in the effluent than influent samples (Table S3). 3.4. Abundance of ARGs in bio-treatment processes

3.3. Identification of ARB and ARGs Considering that WWTPb had a relatively stable performance in removing ARB (Table 1), 16S rRNA gene of 188 randomly selected isolates (94 from influent and 94 from effluent) from this plant in July were cloned and sequenced. In total, eighteen genera were identified in 188 isolates (Tables 3 and S4). Gram-negative (n = 65) and -positive isolates (n = 64) were dominant in the influent and the effluent, respectively (Tables 3 and S3). The mean MAR indices of isolates in these genera ranged from 0.46 to 0.78 and from 0.46 to 0.77, respectively, in the influent and effluent samples. Proteobacteria occurred in 69.1% and 29.8% of the isolates from the influent and effluent, respectively, and for Firmicutes they were 29.8% and 67.0%, respectively. Actinobacteria and Bacteroidetes were also detected. Among the eighteen genera, Aeromonas and Bacillus were detected at a higher frequency as compared to the others. The presence of 13 ARGs (including sul1, sul2, sul3, sulA, tetA, tetB, tetE, tetA, strA, strB; TEM, SHV, CTX-M) were detected in the corresponding antibiotic-resistant isolates from the influent and effluent samples of WWTPb (Tables 4 and S5–6). The ARGs were detected at a higher frequency in ARB from the influent than the effluent samples, except for sulA and CTX-M. The sulphonamide resistance genes sulI, sulII, sulIII and sulA occurred in 36%, 29%,

Among the three WWTPs, the smallest and the largest reduction in MAR index values were detected in WWTPa and WWTPb (Table 1), respectively. Therefore, the 13 ARGs were destined in activated sludge from anaerobic, anoxic and aerobic compartments of these two plants (Tables 5 and S7). In general, abundance of the ARGs increased in aerobic sludge as compared to to anaerobic sludge for the same plant at the two sampling times (Table 5). Compared to the relative abundance of ARGs in the anaerobic compartments of WWTPa, 3 and 10 ARGs increased in the anoxic and aerobic compartments in March and 2 and 10 increased in July, respectively (Table S7). Meanwhile, 7 and 2 ARGs decreased in March and 8 and 2 decreased in July, respectively. In WWTPb, the abundance of 12 and 11 ARGs increased in the anoxic and aerobic compartments, respectively, from the anaerobic compartment in March, while 11 and 10 ARGs increased, respectively in July. However, only 2 ARGs decreased in the aerobic compartment in March. Among the four SUL resistance genes, the relative concentration of sulA was lower than that of other SUL genes (Table S7). The relative level of tetO (0.01–4.89  10 1) was higher than that of tetA (0.11–6.01  10 3), tetB (0.01–24.80  10 4), or tetE (0.39– 80.50  10 3). The level of tetA was higher in the aerobic tank than the anoxic tank, while a reverse trend was observed in the abundance of tetB. The level of strA was similar to that of strB, while the level of CTX-M was generally higher than those of TEM and SHV.

Table 1 Concentrations (log10) and removal rate of the total cultivable heterotrophs and total coliforms in the influents and the final effluents of three WWTPs. Time

WWTPs

Total heterotrophs

Total coliforms

Influent

Effluent

Removal rate (%)

Influent

Effluent

Removal rate (%)

March

WWTPa WWTPb WWTPc

5.87(0.03)* 6.01(0.03)* 6.53(0.02)*

5.26(0.03) 3.59(0.04) 4.81(0.02)

75.45 99.62 98.10

4.68(0.04)* 4.33(0.04)* 4.65(0.28)*

3.18(0.07) 2.59(0.14) 2.70(0.07)

96.84 99.27 98.89

July

WWTPa WWTPb WWTPc

5.69(0.06)* 5.78(0.06)* 6.22(0.08)*

5.07(0.02) 3.01(0.02) 4.11(0.07)

76.01 99.83 99.23

4.07(0.05)* 4.21(0.05)* 4.62(0.59)*

2.28(0.05) 1.59(0.26) 1.60(0.16)

98.37 99.76 99.90

* Indicates the number of total heterotrophs or total coliforms was higher in influent than in effluent of the WWTP (P < 0.01). The value was expressed as mean (S.D.) and the number in parenthesis was the standard deviation.

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Table 2 Percentages (%) of antibiotic resistance bacteria in 200 isolates from influents and effluents of three WWTPs. Antibiotics/MAR index

WWTPa

WWTPb

March

AMP CFZ CFR TET OTC NA LEV GEN STR SMZ VAN CHL MAR index

March

July

Mean

March

July

INF

EFF

INF

EFF

INF

EFF

INF

EFF

INF

EFF

INF

EFF

INF

EFF

Total

96 88 33 64 79 96 30 90 79 82 94 97 0.77

99 88 58 25 79 99 21 88 94 83 96 94 0.77

93 94 35 64 79 79 6 44 43 92 73 98 0.67

85 93 41 56 66 56 3 38 36 82 35 95 0.57

96 88 33 64 79 96 30 90 79 82 94 97 0.77

77 93 30 36 56 93 27 77 50 89 67 90 0.65

93 94 35 64 79 79 6 44 43 92 73 98 0.67

72 93 30 83 27 62 3 47 38 67 62 88 0.56

94 92 61 68 82 100 13 94 84 80 96 100 0.80

99 88 58 25 79 99 21 88 94 83 96 100 0.78

94 95 35 74 69 63 0 34 25 92 75 94 0.63

67 98 25 15 33 63 0 35 28 82 90 100 0.53

94 92 39 66 78 86 14 66 59 87 84 97 0.72

83 92 40 40 57 79 13 62 57 81 74 95 0.64

89 92 40 53 67 82 13 64 58 84 79 96 0.68

Number in ‘italic’ indicates the resistance rates in effluents, which were higher than that in influents in the same plant at the same sample time. INF, influent; EFF, effluents.

Frequency of multi-drug resistance



July

WWTPc

100% 90%

R=11~12

80%

R=10

70% 60%

R=9

50%

R=8

40%

R=7

30%

R=6

20% R=0~5

10% 0%

INF EFF INF EFF INF EFF INF EFF INF EFF INF EFF

March July WWTPa

March July WWTPb

July March WWTPc

Fig. 1. Distribution of multi-drug resistance cultivable heterotrophic isolates in the influent (INF) and effluent (EFF) of three waste water treatment plants (WWTPa, WWTPb and WWTPc). R is the numbers of antibiotics that isolates resist.

4. Discussion 4.1. WWTPs decreased the number and antibiotic resistance of total cultivable heterotrophs The numbers of total heterotrophs were higher in the influent than effluent samples of the three WWTPs (Table 1). Similar results

were obtained in other WWTPs located in different countries or regions (Guardabassi et al., 2002; Reinthaler et al., 2003; Ferreira da Silva et al., 2007; Servais and Passerat, 2009; Zhang et al., 2009; Novo and Manaia, 2010; Gao et al., 2012). Decreases in total coliforms are consistent with previous reports that wastewater treatment resulted in significant decreases in total coliforms (Ferreira da Silva et al., 2007), E. coli concentration (Reinthaler et al., 2003; da Costa et al., 2008) and in the number of faecal indicators (Łuczkiewicz et al., 2010). These data demonstrate that WWTPs have facilitated to remove the bacteria from sewage. Our study showed that the ratio of antibiotic resistant cultivable heterotrophs in waste water decreased after treatments (Table 2). Similar result was also reported in other WWTPs, although some increases in antibiotic resistance were also observed (Garcia et al., 2007). Munir et al. (2011) reported that high concentrations of TET- and SUL-resistant bacteria in the raw sewage of five WWTPs were reduced by several orders of magnitude by the treatments. With the same antibiotics used in this study (Table 2), the resistance rates of the isolates were near to or even higher than those of faecal coliforms from the other WWTPs (Reinthaler et al., 2003; Łuczkiewicz et al., 2010; Huang et al., 2012). For all the three WWTPs, the mean percentages of isolates showing resistance to single antibiotics were lower in the effluent than influent samples (Table 2), with the exception of

Table 3 Occurrence of genera in 188 cultivable heterotrophic isolates from WWTPb. Influent (n = 94)

Effluent (n = 94)

Phylum

Genus

Number

%

MAR index

Number

%

MAR index

c-Proteobacteria

Aeromonas( ) Bacillus(+) Citrobacter( ) Ensifer( ) Enterobacter( ) Escherichia( ) Hydrogenophaga( ) Klebsiella( ) Lysinibacillus(+) Microbacterium(+) Myroides( ) Pannonibacter( ) Proteus( ) Providencia( ) Pseudomonas( ) Serratia( ) Sphingomonas( ) Staphylococcus(+)

46 24 2 / 6 1 / 2 3 1 / 1 5 2 / / / 1

48.94 25.53 2.13 / 6.38 1.06 / 2.13 3.19 1.06 / 1.06 5.32 2.13 / / / 1.06

0.75 0.58 0.66 / 0.78 0.62 / 0.77 0.69 0.46 / 0.54 0.62 0.58 / / / 0.54

14 48 / 2 / / 1 / 14 1 2 / / 5 4 1 1 1

14.89 51.06 / 2.13 / / 1.06 / 14.89 1.06 2.13 / / 5.32 4.26 1.06 1.06 1.06

0.63 0.51 / 0.66 / / 0.54 / 0.5 0.46 0.69 / / 0.51 0.73 0.69 0.77 0.62

Firmicutes

c-Proteobacteria a-Proteobacteria c-Proteobacteria c-Proteobacteria b-Proteobacteria

c-Proteobacteria Firmicutes Actinobacteria Bacteroidetes a-Proteobacteria c-Proteobacteria c-Proteobacteria c-Proteobacteria c-Proteobacteria a-Proteobacteria Firmicutes

/ Indicates that species was not detected in the genus. +, Gram positive isolate;

, Gram negative isolate.

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S. Zhang et al. / Chemosphere 135 (2015) 138–145 Table 4 Percentages of ARGs in the corresponding antibiotic resistance Gram-negative (G ) and -positive (G+) isolates from WWTPb. Antibiotics

Gene

Influent (94)

Effluent (94)

G

G+

Total

G

G+

Total

Sulphonamides

sul-1 sul-2 sul-3 sul-A TARGs TARB

48% 33% 38% 25% 76% 63

8% 19% 4% 4% 35% 26

36% 29% 28% 19% 64% 89

57% 57% 26% 48% 87% 23

11% 11% 2% 6% 21% 47

26% 26% 10% 20% 43% 70

Tetracycline

tetA tetB tetE tetO TARGs TARB

24% 17% 71% 36% 91% 58

7% 7% 48% 85% 93% 27

19% 14% 64% 52% 92% 85

26% 5% 53% 21% 68% 19

7% 3% 17% 67% 83% 30

14% 4% 31% 49% 78% 49

Streptomycin

strA strB TARGs TARB

63% 40% 78% 63

38% 24% 52% 29

55% 35% 70% 92

30% 13% 47% 30

22% 21% 33% 58

25% 18% 38% 88

b-Lactamases

TEM SHV CTX-M TARGs TARB

83% 37% 14% 89% 65

83% 21% 0 86% 29

83% 32% 10% 88% 94

73% 30% 63% 97% 30

67% 24% 10% 69% 58

69% 26% 28% 78% 88

TARGs indicates the percentages of bacteria carrying at least one ARG; TARB indicates the numbers of antibiotic resistance bacteria for the same class of antibiotics. G+, Gram-positive isolate; G Gram-negative isolate.

CFR-resistant isolates, which were slightly more abundant in the effluent than the influent samples. Whether the wastewater treatment process can increase the prevalence of antibiotic-resistant bacteria remains controversial (Zhang et al., 2009). For example, treatment processes positively selected antimicrobial resistance patterns of faecal indicators (Łuczkiewicz et al., 2010) and Acinetobacter (Zhang et al., 2009), while several reports from Guardabassi et al. (2002) suggested that tertiary wastewater treatment did not result in a selection of antimicrobial-resistant bacteria and that the percentages of resistant bacteria in the WWTP were mainly observed to decrease during the treatment processes (Iwane et al., 2001). The conflicting reports might be ascribed to the different target species, which have been exposed to different types and doses of drugs. The MAR index has been widely used to understand the potential resistance of isolates to multiple antibiotics (Servais and Passerat, 2009; Odjadjare et al., 2012) and the higher the multi-resistance level, the higher MAR value. The mean MAR indices of the effluent samples (0.72) were lower than that of the influent samples (0.64) (Table 3), suggesting that WWTPs can reduce the level of antibiotic resistance for total cultivable heterotrophic bacteria. The MAR index of the effluent samples (Table 2) was higher than that of E. coli isolates from environmental water samples (Wang

et al., 2013a) and from the poultry slaughterhouse WWTPs (da Costa et al., 2008). However, abundant MAR heterotrophic bacteria were released into the receiving water although the WWTPs effectively removed heterotrophic bacteria, (Huang et al., 2012). Therefore, sufficient disinfection and appropriate operating conditions should be applied to prevent the spread of antimicrobial-resistant bacteria (Łuczkiewicz et al., 2010; Korzeniewska et al., 2013). Phyla (Firmicutes, proteobacteria, Actinobacteria and Bacteroidetes) identified in this study (Tables 3 and S4) were also found dominated in the activated sludge of WWTPs (Wang et al., 2013b; Yang et al., 2014). It should be noted that Aeromonas (48.94%) were dominant in the influent, but Bacillus (51.06%) were dominant in the effluent samples. Genera as Providencia, Pseudomonas, Sphingomonas and Staphylococcus occurred at lower frequency in the effluent and the mean MAR indices of Aeromonas and Bacillus isolates decreased after treatment in the WWTPs. Hoa et al. (2011) reported the abundances of 19.8% for Aeromonas and 13.2% for Bacillus in 121 SMX-resistant isolates from the Red River delta of northern Vietnam. 4.2. ARGs decreased in cultivable heterotrophs but increased in sludge The antibiotic resistance of these isolates can be ascribed to the intrinsic resistance of the isolates with lower susceptibility to antibiotics (García-León et al., 2014). Sulfamethoxazole resistance is always associated with the protein encoded by the dihydropteroate synthase gene, which has been frequently detected in surface water, agricultural soils, and livestock lagoons (Su et al., 2012; Wang et al., 2013a). The detection rates of suls among the heterotrophic isolates in this study were lower than those of the E. coli isolates from surface water in the same region (Wang et al., 2013a). sul genes were detected in 16 out of 18 genera, while the report of Hoa et al. (2011) showed that 23 out of 25 genera in aquatic environments of northern Vietnam contained sul genes. The frequencies of tetO were higher in Gram-positive genera than Gram-negative genera (Tables S5 and S6), as tetO gene is primarily associated with Gram-positive genera (Luna and Roberts, 1998). However, the detection frequency of strA was lower than that reported by Wang et al. (2013a). The frequency of TEM gene in Aeromonas isolates (Tables S5 and S6) was lower than that reported by Igbinosa and Okoh (2012). Lu et al. (2010) reported that the CTX gene and TEM gene were dominant (22 strains) in ESBL-producing bacteria from an urban river sediment habitat. They also found that most of the ESBL-producing bacteria (Escherichia, Klebsiella, Serratia, and Aeromonas) were dominant in the population. The antibiotic resistance rates and the homogeneity of the removal of organisms belonging to different resistance groups might be strongly influenced by treatment efficiency (Novo and Manaia, 2010). The MAR indices for ESBL-positive isolates were higher in the present study than those for the ESBL-negative isolates

Table 5 Fates of ARGs in anoxic and aerobic sludge compared to that in anaerobic sludge of two WWTPs at two sampling time*. Genes

sul1 sul2 sul3 sulA TEM SHV CTX-M

Increase

Decrease

Genes

Anoxic

Aerobic

Anoxic

Aerobic

0 2 2 3 3 2 2

2 4 3 2 4 3 3

3 1 2 0 0 2 1

1 0 2 2 0 1 0

tetA tetB tetE tetO strA strB

Increase

Decrease

Anoxic

Aerobic

Anoxic

Aerobic

2 1 3 3 2 1

4 3 4 2 4 3

1 0 1 1 0 2

0 0 0 2 0 0

* The figures indicate the increase or decrease of ARGs in 4 sets of samples from aerobic compartments compared to those from anaerobic compartments (for details see supplementary materials Table S7).

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(0.45–0.63) (Korzeniewska et al., 2013). As many of these isolates are opportunistic pathogens or indigenous to aquatic environments as important opportunistic pathogens (Aminov, 2009; Janda and Abbott, 2010), they have the potential to develop and spread antibiotic resistance in the environment, posing a threat to public health (Shangkuan et al., 2000; Arias et al., 2010) and/ or altering the antibiotic resistance and microbial diversity of the bacterial community in the receiving environment (Suzuki et al., 2008; Hoa et al., 2011). The WWTPs with combined anaerobic, anoxic and aerobic processes have been widely employed to treat wastewaters. The relative abundance of ARGs (normalised to the copy numbers of 16S rRNA genes) was assayed by an RT-PCR method to analyse changes in their abundance during the anaerobic, anoxic and aerobic treatment processes in this study. In general, the concentrations of ARGs in sludge increased with the increase of treatment process (Tables 5 and S7). Morozzi et al. (1988) found that, during aerobic treatment of urban sewage, there was a tendency for the percentage of ARB to increase. A recent report showed that aerobic and anaerobic sludge differed with respect to the abundance of different types of ARGs and the bacterial community structure and composition (Wang et al., 2013b). However, the detection frequency of ARGs decreased generally in cultivable heterotroph in the effluent, as compared to the influent (Tables S5 and S6). This suggests that abundant uncultivable and inactive or dead bacteria may harbour high level of ARGs. For example, Börjesson et al. (2010) reported a reduction in tetA and tetB gene concentrations in sewage after the wastewater treatment process. They ascribed the reduction of tetA and tetB partly to the sedimentation process. Our data showed that different WWTPs had different efficiency in removing ARGs and ARBs (Tables 2 and S7). These differences can be ascribed to different treatment technology (sequence of compartments and performance parameters (Table S1) such as pH and hydraulic/sludge retention time) and the source of sewage (Morozzi et al., 1988; Novo and Manaia, 2010; Moura et al., 2012; Chen and Zhang, 2013). For example, the availability and types of nutrients and oxygen level always varied in anaerobic, anoxic and aerobic compartments of WWTPs to adjust the bacteria community for optimizing nutrient removal efficiency (Table S1). Additionally, WWTPs receive not only ARBs but also abundant antibiotics (and heavy metals), which contribute to the development and spread of antibiotic resistance (Stepanauskas et al., 2006). For example, three b-lactamase genes detected in this study were also frequently detected in isolates from hospital effluent and municipal sewage (Korzeniewska et al., 2013). Therefore, wastewater types and sources were also key factors that affect the performance of WWTPs with respect to ARBs and ARGs removal. Taken together, the development and spread of ARB and ARGs are complex processes in WWTPs and wastewater treatment systems remain the important reservoirs for various ARGs. Therefore, the discharge of sludge and even treated sewage may contribute to release of ARGs and ARB to the environment (Guardabassi et al., 2002; Auerbach et al., 2007; Gao et al., 2012; Moura et al., 2012).

5. Conclusions The spread of ARB and ARGs in the treatment processes of WWTP remains a problem, and research is needed to understand the fate of antibiotic-resistant cultivable heterotrophic bacteria and antibiotic resistance genes in wastewaters. The results from the present study first time identified and characterized the phenotypes and genotypes of antibiotic-resistant heterotrophic bacteria in the influent and effluent samples from different WWTPs and determined the abundance of ARGs in anoxic, anaerobic and aerobic WWTP compartments. Our results indicated that there is a

critical need to more fully understand the processes that control the development of antibiotic resistance in WWPTs with the increased use of recycled wastewater and biosolids for irrigation and soil management, respectively. All the WWTPs displayed adequate performance in the removal of total cultivable heterotrophic bacteria, including faecal coliforms. The concentrations of >10drug-resistant bacteria and the multi-antibiotic resistance indices were generally reduced after wastewater treatment. 16S rRNA gene sequence analysis revealed eighteen genera among the 188 isolates from the influent and effluent samples of WWTPb, and only six genera (Aeromonas, Bacillus, Lysinibacillus, Microbacterium, Providencia, and Staphylococcus) were detected in both the influent and effluent samples. The MAR indices of Gram-negative isolates were higher than those of Gram-positive isolates. Thirteen ARGs belonging to the tetracycline, sulphonamide, streptomycin and blactam resistance classes were detected at a higher frequency in ARB from the influent than effluent, with the exception of CTX-M. These results suggest that WWTP systems contribute to a decrease in the antibiotic resistance of total heterotrophic isolates in sewage by altering the structure of the bacterial community. In general, the abundance of 13 ARGs was higher in aerobic compartments than in the anoxic tank, suggesting that the ARG abundance increased in activated sludge after WWTP processes. In this study, we only investigated the fate of ARB and ARGs in three WWTPs with anoxic, anaerobic and aerobic compartments. In fact, there are other WWTPs with different technologies in the world, and abundant heterotrophic bacteria and ARGs exist in these WWTPs. Moreover, the development and spread of ARGs are sophisticated. Therefore, further work should be performed in combination with meta-genomic methods in future. Acknowledgments This study was, in part, supported by Grants from the National Natural Science Foundation of China (Grant No. 51379063), Jiangsu Natural Science Foundation (BK2012413), Science Fund for Creative Research Group of the National Natural Science Foundation of China (51421006) , and Innovation Project from Ministry of Education of China (IRT13061) and PAPD. Appendix A. Supplementary material Supplementary data associated with this article can be found, in the online version, at http://dx.doi.org/10.1016/j.chemosphere. 2015.04.001. References Aminov, R.I., 2009. The role of antibiotics and antibiotic resistance in nature. Environ. Microbiol. 11, 2970–2988. Aminov, R.I., Garrigues-Jeanjean, N., Mackie, R.I., 2001. Molecular ecology of tetracycline resistance: development and validation of primers for detection of tetracycline resistance genes encoding ribosomal protection proteins. Appl. Environ. Microbiol. 67, 22–32. Arias, A., Seral, C., Gude, M.J., Castillo, F.J., 2010. Molecular mechanisms of quinolone resistance in clinical isolates of Aeromonas caviae and Aeromonas veronii bv. sobria. Int. Microbiol. 13, 135–141. Auerbach, E.A., Seyfried, E.E., McMahon, K.D., 2007. Tetracycline resistance genes in activated sludge wastewater treatment plants. Water Res. 41, 1143–1151. Baquero, F., Martinez, J.L., Canton, R., 2008. Antibiotics and antibiotic resistance in water environments. Curr. Opin. Biotechnol. 19, 260–265. Börjesson, S., Mattsson, A., Lindgren, P.E., 2010. Genes encoding tetracycline resistance in a full-scale municipal wastewater treatment plant investigated during one year. J. Water Health 82, 247–256. Bouki, C., Venieri, D., Diamadopoulos, E., 2013. Detection and fate of antibiotic resistant bacteria in wastewater treatment plants: a review. Ecotoxicol. Environ. Saf. 91, 1–9. Chen, H., Zhang, M., 2013. Occurrence and removal of antibiotic resistance genes in municipal wastewater and rural domestic sewage treatment systems in eastern China. Environ. Int. 55, 9–14.

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Fate of antibiotic resistant cultivable heterotrophic bacteria and antibiotic resistance genes in wastewater treatment processes.

Antibiotic resistant bacteria (ARB) and antibiotic resistance genes (ARGs) are emerging contaminants of environmental concern. Heterotrophic bacteria ...
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