The basic function of wastewater treatment plants (WWTP) is to remove of debris, high organic loads and pathogens from wastewater, containing mainly household sewage and some industrial wastewater, before discharging into environmental receptors (water streams/rivers, lakes, sea). However, human activities have resulted into an increase release of antibiotics, antibiotic resistance genes (ARGs) and antibiotic resistant bacteria (ARB) into environment. In WWTPs, pharmaceuticals, ARB and ARGs released through several routes, i.e., hospitals with urine and faeces from patients and, similarly, via people taking antibiotics at home, livestock and poultry with excrements from animals treated with antibiotics and environmental bacteria are combined. Co-occurrence of high density of faecal bacteria and antibiotics at subinhibitory concentration is a key factor for the acquisition and proliferation of antibiotic resistance among bacteria. Researchers reported links between antimicrobial residues in sewage or environment, bacterial community structure and composition and antibiotic resistance (Zhang et al., 2009; Jury et al., 2011; Novo et al., 2013). Thus, WWTPs are considered resistance hotspots, providing adequate conducive conditions for the establishment and propagation of antibiotic resistance through biological processes of selection and gene transfer (Jury et al., 2011; Rizzo et al., 2013; Williams et al., 2016). Additionally, the preferential elimination of susceptible organisms rather than their resistant counterparts may have an important contribute for resistance increase (Figueira et al., 2011).
Wastewater treatment is able to reduce the total bacterial load, but evidence for the selection of extremely multiresistant strains and accumulation of resistance genes was observed (Czekalski et al., 2012). Bacteria residing in different compartments of the same WWTP were demonstrated to harbor various plasmid-borne resistance determinants, representing all common classes (Szczepanowski et al., 2009). Additionally, WWTPs effluents have been demonstrated to contain ARB and ARGs, at high prevalence rates, that are released into environment (de Silva et al., 2006; Luczkiewicz et al., 2010; Vaz-Moreira et al., 2014; Manaia et al., 2016; Sato and Urase, 2016). However, large studies are required in order to set a framework and develop programs to improve the wastewater treatment technologies performance for the removal of ARB and thus combating antibiotic resistance.
Diversity and abundance of antibiotic resistance in WWTP
WWTP are interfaces between different environments and, therefore, provide an opportunity for mobile elements (including resistance) to mix between pathogens, opportunistic pathogens and environmental bacteria (Szczepanowski et al., 2009). The presence of antibiotics in sewage selects for drug resistance markers that are able to spread through the microbial community and, as a result, ARB can potentially disseminate their ARGs widely among members of the endogenous microbial community. Additionally, the increase of ARBs and ARG into environment may enhance the horizontal gene transfer (HGT) and the acquisition of drug resistance by environmental bacteria (Lupan et al., 2017). The sludge products of urban and rural wastewater treatment plants are increasingly used to fertilize agricultural crops, dispersing unknown amounts of ARG, ARB and antibiotics that withstand standard sewage treatment.
Various approaches are used to investigate the diversity and abundance (per sample mass or volume) or prevalence (per total bacteria) of antibiotic resistance in WWTP (Manaia et al., 2018). However, most of the studies focus on a few specific bacterial genera as indicators of overall antibiotic resistance, and thus do not reflect the dynamic of microbial communities (Jury et al., 2011). The analysis methods are classified as culture-dependent and culture-independent, based on the use of cultivation methods and analysis of nucleic acids (DNA or RNA), respectively. Phenotypic investigation of antibiotic resistance will not offer a comprehensive view of resistance levels because less than 10% of wastewater bacteria may be cultivated in laboratory conditions (Jury et al., 2011; Vaz-Moreira et al., 2013; Manaia et al., 2018). The molecular-based methods have the advantages that they yield important information about the largest majority of environmental bacteria and the amount of genes encoding antibiotic resistance. In order to explore the diversity of resistomes in environmental samples, an integrated approach combining cultivation and molecular based methods may represent the best choice (Batt et al., 2007; Port et al., 2014; Yang et al., 2014; Li et al., 2015).
Conventional culture‑based methods for assessing resistance associated with WWTP
ARB, and even resistant pathogens have been isolated from WWTP and their effluents using cultivation methods (Figueira et al., 2011; Rizzo et al., 2013). Membrane filtration method (ISO 9308-1:2000 and ISO 7899-2:2000) is used to isolate and enumerate bacteria belonging to different groups, that are present in selected niches within the urban water cycle (Ferreira da Silva et al., 2006; Vaz-Moreira et al., 2011). After filtration of adequate volume and decimal dilutions of wastewater samples, the filters are placed onto media for isolation and enumeration of bacteria – R2 agar, Plate Count Agar for heterotrophs (Ferreira da Silva et al., 2005; Czekalski et al., 2012; Novo et al., 2013); m-Faecal Coliforms agar, EMB, Lactose TTC for Enterobacteriaceae (Prado et al., 2007; Czekalski et al., 2012; Novo et al., 2013); TBX for E. coli (Blaak et al., 2015; Franz et al., 2015); m-Enterococcus agar, Bile Esculin Agar, Enterococcus selective agar for enterococci (Ferreira da Silva et al., 2006; Luczkiewicz et al., 2010) Pseudomonas Isolation Agar for Pseudomonas spp. (Czekalski et al., 2012); Agar LAM for Acinetobacter (Zhang et al., 2009). To determine the ARB presence in wastewater samples, the filters are placed onto the same media or nutrient agar supplemented with different antibiotics for the determination of corresponding antibiotic resistant subpopulations (Watkinson et al., 2007; Novo et al., 2013; Yuan et al., 2015) or onto selective agars: ChromID ESBL agar (bioMérieux, Marcy l’Etoile, France), Drigalski agar or Tryptone Bile X-glucuronide medium supplemented with antibiotics (Franz et al., 2015; Drieux et al., 2016). The percentage of resistance for each antibiotic or antibiotic class corresponds to the ratio of CFU ml-1 on the culture medium with and without antibiotic (Novo and Manaia, 2010; Novo et al., 2013). There are studies that addressed the issue in terms of the presence or absence of antimicrobial-resistant bacteria (Ferreira da Silva et al., 2007; Prado et al., 2008). However, quantitative microbial risk assessment is important in efforts to assess the potential risk and to evaluate the impact of specific elements of the effluent treatment process in removing ARB. Galvin et al. (2010) reported a modification of the most probable number (MPN) method for enumeration of antimicrobial-resistant E. coli bacteria in wastewater samples.
The use of cultivation based methods to assess the anthropogenic impact has revealed a high abundance of ARB in WWTP effluents that are constantly discharged into environment (Huang et al., 2011; Vaz-Moreira et al., 2014; Franz et al., 2015; Blaak et al., 2015; Manaia et al. 2016; Drieux et al., 2016). The concentration of cefotaxime-resistant E. coli strains reported by Galvin et al. in the effluents of an Irish hospital was of 105 UFC/L (Drieux et al., 2016). Blaak et al. (2015) determined median ESBL-concentrations of 2×107, 8.2×105, 1.5×103 in health care institutions WWTP, WWTP influents and WWTP effluents, respectively. In another study, it was determined that ciprofloxacin-resistant coliforms are released from WWTP effluents with an average of 1.6x1013 ± 3.7x1013 per day into the environment (Vaz-Moreira et al., 2014). Huang et al. (2011) reported that the concentrations of penicillin-, ampicillin-, cephalothin-, chloramphenicol-, tetracycline, and rifampicin-resistant bacteria in the secondary effluents of a municipal WWTP were: 1.5×104–1.9×105; 1.2×104–1.5×105; 8.9×103–1.9×105; 2.6×104–2×105; 840–6.1×103 and 310–6.1×104 CFU/mL, respectively. These data are indicating high rates of ARB in environment, which constitute a serious threat posed on human health.
The characterization of ARB isolates concerning their metabolic features, antibiotic susceptibility patterns, ARGs and mobile genetic elements contained is important for epidemiological studies (Guardabassi et al., 2002; Vaz-Moreira et al., 2014; Kaplan et al., 2015; Varela et al., 2015; Hembach et al., 2017). In a study that analyzed E. coli isolates from Dutch surface water (n=113), and WWTP including healthcare institution plants, airport and municipal (n=33), the antibiotic resistance profile of surface water isolates indicated that 26% were resistant to at least one antibiotic and 11% were multiple drug resistant. The highest level of resistance was detected in healthcare institution plants, with 62% of the E. coli isolates multidrug resistance. A total of 20% of the municipal WWTP effluent isolates were multidrug resistant, indicating that wastewater treatment plants are spreading antibiotic resistance in the environment (Blaak et al., 2015). The antibiotic susceptibility testing of E. coli isolates (1326 and 451 isolates from municipal and hospital wastewaters, respectively) showed that 34% of the isolates from municipal wastewater and 55% of the isolates from hospital wastewater were resistant to at least one antibiotic (Kwak et al., 2015). The characterization of 85 enterococcal isolates collected from WWTP and surface water samples showed that they harbor ARGs that included: aph(3’)-IIIa (29%), ant(6)-la (8%), erm(B) (44%), tet(M) (8%) (Ben Said et al., 2015). Zhang et al. (2015) reported that AMR heterotrophic bacteria isolated from influent and effluent water from three WWTP exhibited between 5% and 64% resistance to more than nine antibiotics and those from influents showed higher isolates portions.
Culture independent study of antibiotic resistance associated with WWTP
The application of standard susceptibility methods, based on the cultivation of the microorganisms, cannot provide a complete view about the prevalence and abundance of antibiotic resistance in environmental communities. For this reason, culture-independent methods to characterize antibiotic resistance in environmental samples have been advanced (Bengtsson-Palme et al., 2017). These methods target DNA for ARG monitoring and more rarely RNA for gene expression or proteins in environmental samples. The extraction of nucleic acids is influenced by other components in the sample which may cause an incomplete reaction or completely prevent the chemical reaction from occurring, thereby reducing the recovery of the target DNA (Terrat et al., 2012; Thomas et al., 2012; Knauth et al., 2013; McCarthy et al., 2015; Li et al., 2018). Additionally, other factors – i.e., DNA extraction protocols, sample storage conditions – may influence the results before the application of various sequencing techniques (Bengtsson-Palme et al., 2017). A recent study performed a comparison of sample pretreatment and DNA extraction methods (Li et al., 2017). Three DNA extraction kits were evaluated: FastDNA® Spin Kit for Soil, PowerSoil® DNA Isolation Kit, and ZR Fecal DNA MiniPrep™. The results indicated similar trends when comparing the influent, activated sludge, and effluent samples obtained from two WWTPs located on opposite sides of the globe. However, considering both yield and quality and the diversity of ARGs and taxa from the extracted DNA, plus the reproducibility of the extraction, the FastDNA® Spin Kit for Soil exhibited superior performance for all sample locations (Li et al., 2018).
Two methodologies have been used to describe the diversity and measure the abundance of ARGs in wastewater samples: quantitative PCR and metagenomics. These methods avoid culture bias and provide direct access to the total DNA in a sample.
Real-time PCR technology has allowed the advancement of molecular diagnostics toward a high-throughput technology. The qPCR assay is based on the measurement of the fluorescence produced by a reporter molecule that binds to the double-stranded DNA which increases as the reaction proceeds. There are various detection chemistry; among them, SYBR Green I and TaqMan assays are the most widely used methods in current environmental microbial ecology studies. The absolute quantification of target microbial populations is the primary purpose of qPCR in studies of environmental systems, including wastewater treatment processes. The abundances of ARGs are expressed as the amount ratios of target genes to the reference gene, normally the 16S rRNA gene. Studies have demonstrated that ARG polluted waters released from WWTP enhance the abundance of ARG in sediments downstream from WWTP (Pruden et al., 2012; Czekalski et al., 2014; Amos et al., 2015). Tetracycline and sulfonamide resistance genes were detected in sediments of Lake Geneva (Switzerland), near the WWTP effluent discharge (Czekalski et al., 2014). qPCR assay revealed ARGs and transposase genes related to HGT in samples collected from a large WWTP in Helsinki, Finland. A higher abundance of ARGs and transposase gene was detected in raw sewage compared with effluents and dried sludge leaving the treatment plant (Karkaman et al., 2016). This might be explained by the long use in therapy of certain antibiotics – i.e., sulfonamides and tetracyclines. The molecular analysis of sewage and influents showed a high level of sul and tet genes (Pruden et al., 2006; Lupan et al., 2017) and in livestock waste sludge (Zhang et al., 2013). Lupan et al. (2017) investigated the removal efficiency of newly modernized plant and found that, even though there is a significant reduction of ARB density, the plant is contributing to an increase in ARB up to 10 km downstream of its discharge. The pollution with heavy metals represent an important factor for ARG co-selection (Berg et al., 2010) due to their co-occurrence on the same mobile genetic element (Baker-Austin et al., 2006; Knapp et al., 2011).
The qPCR assay allows the measurement of all DNA, including the DNA from live cells, viable but not cultivable cells, dead cells, and the DNA that is found outside of a cell and in the environment (free DNA), which leads to an overestimation of the active population (Nocker and Camper, 2006). An alternative is the use of DNA-intercalating dyes – for example, ethidium monoazide (EMA) and propidium monoazide (PMA), which selectively removes nonviable DNA (Nogva et al., 2003; Cawthorn and Witthuhn, 2008). qPCR and other nucleotide sequence based methods require known sequences for the detection of target microorganisms. Therefore, it cannot detect uncharacterized sequences, which constitutes a limitation in studying unknown microbial diversity (Lim et al., 2011). Additionally, sequence data about the target populations are very important for designing highly specific primers and probes that minimize the potential for false-positive or false-negative results. Additionally, it does not directly confirm the functionality of the target within a viable host cell; such information may be indirectly determined through combination with other methods and adequate experimental design. All these restrains should be considered when results are to be analyzed and interpreted.
High-throughput sequencing has widened the scope of microbial analysis of environmentally derived samples (He et al., 2017) and brought classical environmental studies to another level. qPCR is a targeted analysis, meaning that it detects only the target. Metagenomics overcomes such limitation through direct genetic analysis of genomes contained within an environmental sample (Schmieder and Edwards, 2012). The first metagenomic studies on microbiota of environmental samples pointed out that 80% of the bacteria identified by metagenomics or by 16S ribosomal RNA (rRNA) pyrosequencing genes had not yet been cultured (Venter et al., 2004).
Metagenomic studies provide a community biodiversity profile that can be further correlated with functional composition analysis of known and unknown organism lineages (i.e., genera or taxa) (Oulas et al., 2015). The results are interpreted as specific versus total reads number, or versus the number of reads of reference gene, most commonly 16 rRNA gene (Parsley et al., 2010; Schmieder and Edwards, 2012; Thomas et al., 2012; Zhang et al., 2012; Tian et al., 2016; Bengtsson-Palme et al., 2017; Louca et al., 2018). The corrections for 16S gene copy numbers in microbiome surveys are not indicated to be performed by default. The OTUs should be sufficiently closely related to genomes sequenced, so that community profiles remain interpretable and comparable between studies (Louca et al., 2018). Reference-based assessment cannot account for unknown species and could result in overestimates into the abundances of known species. Metagenomic assembly may not detect rare species and overestimates the abundance of species (Nayfach and Pollard, 2016). Metagenomic studies are also dependent on the existence of databases for data sets. Currently, such computational tools are becoming more efficient and elaborated as the technology of sequencing improves (Oulas et al., 2015; Breitwieser and Salzberg, 2017).
A number of technical approaches have been used to characterize the phenotypic and genotypic antibiotic resistance in WWTPs. The overall results are indicating that WWTPs are not able to efficiently remove ARB and ARG during treatments. Further studies are required to accurately determine the outcome of ARB and ARG in WWTPs and to establish the impact of their abundance in WWTP effluents on human and environmental health.
The financial support of the ERANET-JPI-EC-AMR -AWARE-WWTP and PN-III-P4-ID-PCCF-2016-0114 is gratefully acknowledged.