Antibiotics 2013, 2, 367-399; doi:10.3390/antibiotics2030367 OPEN ACCESS antibiotics ISSN 2079-6382 www.mdpi.com/journal/antibiotics Review Environmental and Public Health Implications of Water Reuse: Antibiotics, Antibiotic Resistant Bacteria, and Antibiotic Resistance Genes Pei-Ying Hong 1,*, Nada Al-Jassim 1, Mohd Ikram Ansari 1 and Roderick I. Mackie 2,3,* 1 King Abdullah University of Science and Technology (KAUST), Environmental Science and Engineering, Water Desalination and Reuse Center, Thuwal 23955-6900, Saudi Arabia; E-Mails: [email protected] (N.A.-J.); [email protected] (M.I.A.) 2 Department of Animal Sciences, University of Illinois at Urbana-Champaign, Urbana, IL 61801, USA 3 Institute of Genomic Biology, University of Illinois at Urbana-Champaign, Urbana, IL 61801, USA * Authors to whom correspondence should be addressed; E-Mails: [email protected] (P.-Y.H.); [email protected] (R.I.M.); Tel.: +966-02-808-2218 (P.-Y.H.); +1-217-244-2526 (R.I.M.). Received: 17 June 2013; in revised form: 19 July 2013/ Accepted: 24 July 2013/ Published: 31 July 2013 Abstract: Water scarcity is a global problem, and is particularly acute in certain regions like Africa, the Middle East, as well as the western states of America. A breakdown on water usage revealed that 70% of freshwater supplies are used for agricultural irrigation. The use of reclaimed water as an alternative water source for agricultural irrigation would greatly alleviate the demand on freshwater sources. This paradigm shift is gaining momentum in several water scarce countries like Saudi Arabia. However, microbial problems associated with reclaimed water may hinder the use of reclaimed water for agricultural irrigation. Of particular concern is that the occurrence of antibiotic residues in the reclaimed water can select for antibiotic resistance genes among the microbial community. Antibiotic resistance genes can be associated with mobile genetic elements, which in turn allow a promiscuous transfer of resistance traits from one bacterium to another. Together with the pathogens that are present in the reclaimed water, antibiotic resistant bacteria can potentially exchange mobile genetic elements to create the ―perfect microbial storm‖. Given the significance of this issue, a deeper understanding of the occurrence of antibiotics in reclaimed water, and their potential influence on the selection of resistant microorganisms would be essential. In this review paper, we collated literature over the past two decades to determine the occurrence of antibiotics in municipal wastewater and livestock manure. We then discuss how these antibiotic resistant bacteria may impose a potential microbial risk to the Antibiotics 2013, 2 368 environment and public health, and the knowledge gaps that would have to be addressed in future studies. Overall, the collation of the literature in wastewater treatment and agriculture serves to frame and identify potential concerns with respect to antibiotics, antibiotic resistant bacteria, and antibiotic resistance genes in reclaimed water. Keywords: antibiotics; water reuse; antibiotic resistant bacteria; municipal wastewater; livestock manure; manure-applied soil 1. Introduction Water scarcity is a global issue, and is a result of economic and physical constraints. A survey by the International Water Management Institute listed certain regions like Africa, the Middle East, as well as the western states of America as water stressed areas [1]. A breakdown on the freshwater consumptive use in the United States (U.S.) showed that agricultural irrigation accounts for up to 81% of the total daily usage, while consumptive usage by domestic households only accounts for 6% of the total daily usage [2]. Given that applications like agricultural and landscaping irrigation do not require high quality water supply, water reuse has become an attractive option for conserving and extending available water supply. Reclaimed water is technically defined as municipal wastewater that has gone through various treatment processes, and should only be used when the treated water quality falls in line with specific water quality criteria. For instance, the guidelines of U.S. Environmental Protection Agency (USEPA) stated that municipal wastewater would need to undergo secondary and/or tertiary treatment to achieve a considerable reduction in the organic and inorganic constituents, as measured based on biochemical oxygen demand (BOD) and chemical oxygen demand (COD). In addition, appropriate treatment has to be utilized to achieve no detectable fecal coliform or less than 200 fecal coliforms/100 mL in the reclaimed water prior to surface irrigation on any food crop or orchards and vineyards, respectively [3]. As municipal wastewater is a relatively stable and reliable flow of water that is rich in nitrogen and phosphorus content arising from the fecal contents, it is often deemed as an attractive source of water for agricultural irrigation. Livestock production farms in U.S. are no exception as they generally rely on animal manure as a nitrogen and phosphorus rich fertilizer for their agricultural crops. In most instances, the animal manure is first retained in a manure pit or lagoon over a period of time, before application on the agricultural field. Application can be applied by spraying over crops or through a direct injection into the soil at a 20 cm depth [4]. Although the use of municipal wastewater as an alternative water source is a highly attractive option, there is a need to understand the microbial risks arising from antibiotic residues, the antibiotic resistant bacteria, and their associated resistance genes. This is particularly complicated when antibiotics are increasingly consumed for disease treatment in humans and livestock, and as prophylaxis and growth promoters for the latter. The European Surveillance of Antimicrobial Consumption (ESAC) project collected data on antibiotic use for the period 1997–2001, and determined that the median national hospital antibiotic consumption in Europe was 2.1 DDD/1,000 inhabitants/day [5]. Defined daily dose (DDD) is the assumed average maintenance dose per day for a drug used for its main indication in Antibiotics 2013, 2 369 adults [6]. Assuming an average dose of 750 mg for most of the examined antibiotics, that would amount to approximately 1,067 kg of antibiotics consumed per day in Europe. Surprisingly, hospital care consumption only accounts up to 17.8% of the total antibiotic consumption in Europe [5], while majority of the antibiotics are consumed in normal households. In the same manner, livestock production utilizes equal, if not more antibiotics than in the human population. The Union of Concerned Scientists reported that approximately 11 million kg of antibiotics were used non-therapeutically in the swine, poultry and cattle production industries [7]. This accounts for more than 50% of the total antibiotics consumption in U.S. per annum [7]. The rampant use of antibiotics among both human and animal populations clearly suggests that large amounts of antibiotics would end up in municipal WWTPs and in animal manure. Indeed, environmental reservoirs are increasingly viewed as one of the major hotspots for various microorganisms to gain antibiotic resistance. As such, the presence of antibiotics would have to be examined if wastewater is to be reused. In the following review, we seek to provide a comprehensive overview on the types and concentration of antibiotics that are detected in municipal wastewater and livestock manure. We then collate the abundant, various antibiotic resistance genes that are present in municipal wastewaters and treated effluent, as well as their abundance in animal waste lagoon and manure-applied soils. Finally, we discuss the significance of these contaminants pertaining to the use of reclaimed water to alleviate water scarcity. 2. Antibiotics in Municipal Effluent In 2002, an estimated figure on the antibiotics consumed annually worldwide was between 100 million and 200 million kg of antibiotics [8]. A collation on the antibiotic prescription in a Portugese medical study revealed that the most frequently prescribed antibiotics were penicillin (47%), macrolides (16%), quinolones (15%), cephalosporins (12%), sulfonamides (5%) and tetracyclines (2%) [9,10]. In British Columbia, a similar trend in the usage of different classes of antibiotics was also observed. Beta-lactams like penicillin and cephalosporins were most commonly consumed in British Columbia, followed by tetracyclines, macrolides, sulfonamides and fluoroquinolones [11]. The mechanisms of each different class of antibiotics have already been extensively reviewed, and will not be included here. Readers can refer to the review paper by Kohanski to understand the mechanisms underlying the antibiotics mode of action against bacteria [12]. Most of these antibiotics are excreted from the human body, and are excreted as the parent compound in the feces or urine, which in turn ends up in the municipal wastewater treatment plant. In a conventional municipal wastewater treatment process, municipal wastewater is first screened and clarified to remove large particulates and suspended particulates, respectively, prior to the biological treatment process in activated sludge. The treated water after the biological process is then pumped into a secondary clarifier to further remove suspended particulates and organic constituents. Typically, the treated water at this point is termed as the secondary treated effluent, and may undergo post-treatment processes (e.g., sand filtration, membrane filtration or disinfection). The tertiary treated water then gets reused for various purposes such as agricultural irrigation. Based on the treatment schematics, the quality of the effluent is generally anticipated to meet the stipulated guidelines of 5–30 mg/L of BOD [3], and is appropriate for irrigation on food and nonfood crops. Antibiotics 2013, 2 370 With the increasing use of municipal wastewater for various types of reclaimed water purposes, guidelines are also formulated to provide guidance [13] on the minimal achievable quality required for the reclaimed water. However, these guidelines do not include the minimal concentration of antibiotics residues and abundance of resistant bacteria that are allowed in reclaimed water. It is gradually recognized that antibiotics are emerging contaminants that can result in a ―perfect microbial storm‖, defined as a phenomenon where novel microbial threats emerge with elevated frequency and that can create an environment that allows infectious diseases to emerge and become rooted in society [14]. Clearly, the current guideline is not updated with emerging contaminants such as antibiotics residues and antibiotics resistant bacteria. The removal efficiency of various types of antibiotics was recently reviewed [15]. Generally, removal is achieved via chemical treatment and/or bio-adsorption onto particulates and subsequently physical separation from municipal wastewater after sedimentation in the screening chambers and clarifiers. Therefore, the removal efficiency is highly dependent on the hydrophobicity and sorption capability of the antibiotics. In contrast, biological degradation is deemed to be relatively less effective in removing antibiotics from the bulk of the municipal wastewater. Based on this conventional treatment scheme, the abundance of antimicrobials in municipal wastewater treatment in Croatia was found to range from 2–20 μg/L [16]. Antibiotics like sulfapyridine (sulfonamides), azithromycin (macrolide) and erythromycin (macrolide) were not removed effectively. In particular, trimethoprim (dihydrofolate reductase inhibitor) was not sufficiently removed by the treatment process. This low removal efficiency can perhaps be explained by the low sorption potential of most sulfonamides (i.e., logK < 2.5) and ow medium sorption potential for macrolides (i.e., 2.5 < logK < 4) [16,17]. However in certain instances, ow high removal rates of sulfonamides like sulfamethoxazole and quinolones like norfloxacin and ciprofloxacin can still be achieved. Nevertheless, the detected abundance of these antibiotics residues in the secondary effluent still remained high, ranging at concentration of 119–544 ng/L, 24–175 ng/L and 11–168 ng/L, respectively, in the secondary effluent [16]. Alternatively, coupling the conventional treatment process with additional tertiary treatment like membrane filtration can further enhance the removal efficiency of antibiotics. The membrane filtration process achieves liquid-solid separation based on the pore sizes of the attached membrane. Given the small molecular size of antibiotics, microfiltration and ultrafiltration membranes that are connected to the bioreactors would not be able to provide the appropriate size exclusion to remove antibiotics from the municipal wastewater. Instead, a nanofiltration or reverse osmosis membrane would be more appropriate for the removal of antibiotics [18], although a full removal rate does not seem to be attainable for certain antibiotics even when a reverse osmosis membrane is used [19]. Moreover, a membrane bioreactor connected to nanofiltration or reverse osmosis membranes would require a higher operating pressure and therefore incur a higher energy landscape and operating costs. Considering that biofouling and organic fouling are both ubiquitous events on the membrane, these biofilm matrixes can be viewed as an additional barrier that will serve to improve the antibiotic removal efficiency. As such, most of the existing treatment plants which opt for membrane filtration as an additional tertiary treatment step would use microfiltration and ultrafiltration membranes to achieve a reasonably good effluent quality in consideration of the cost and benefits. An additional advantage of the membrane bioreactor is the provision of high solid retention time within the bioreactor. Retentate or reject stream contains antibiotics which did not pass through the Antibiotics 2013, 2 371 membrane, and would be recycled back to the bioreactor. Membrane filtration therefore provides a longer retention time for the antibiotics, and would provide more reaction time for the subsequent breakdown and hydrolysis of the antibiotics through biological degradation. To illustrate, the removal efficiency of certain antibiotics like trimethoprim and macrolides (e.g., erythromycin and clarithromycin) were significantly reduced by up to 90% at a solid retention time period of 60–80 days [20]. Anaerobic membrane processes would also be beneficial for its use to obtain reclaimed water suitable for agricultural irrigation, as nitrogen and phosphorus are retained during an anaerobic biological process while a conventional aerobic municipal wastewater treatment process would remove up to 70% of the nitrogen and 50% of the phosphorus content [21]. Furthermore, anaerobic membrane processes demonstrate the potential to create an energy neutral or positive landscape. However, there is a need to determine the persistence of pathogens and antibiotic resistance genes in the municipal water after going through anaerobic treatment processes. The persistence of plasmids in the different stages of municipal wastewater treatment plant was recently examined [22,23], and contradictory observations were reported. Merlin et al. [22] determined that the copy number of pB10 plasmids to DH5 chromosomal DNA increased over time under anaerobic conditions, and suggest an increase in the occurrence of plasmid transfer. Rysz and coworkers [23], however, noted that a higher rate of antibiotic resistance gene loss was observed in E. coli under anaerobic fermentation conditions than under aerobic conditions. 3. Antibiotic Resistant Bacteria and Associated Genes in Municipal Effluent It is known that the human gut contains bacteria that are intrinsically resistant to antibiotics [24], and thus, many of the bacteria that enter into the treatment system may already be resistant to antibiotics. Furthermore, the existing municipal wastewater treatment design is unable to remove antibiotic resistant bacteria and their associated resistance genes entirely. The activated sludge is therefore exposed to antibiotic residues that may impose a selection for resistant bacteria. The confluence of gut-associated antibiotic resistant bacteria, remnant antibiotic residues and a rich diversity of activated sludge microbial consortium would suggest that the biological treatment process provides favorable environment for mobile genetic elements to be transferred among the microbial communities [15,25]. Past studies have isolated microorganisms from the activated sludge, and examined them for the presence of mobile genetic elements. These studies have demonstrated that many of the antibiotic resistance genes are associated with mobile genetic elements [26,27], which in turn suggest that mechanisms to transfer the genes from one bacterium to another are present. Mobile genetic elements can be transferred in three modes: (i) transformation, a process by which free DNA is incorporated into a competent cell and brings about genetic change in the recipient; (ii) conjugation, a process of genetic transfer that involves cell-to-cell contact; and (iii) transduction, a process by which DNA is transferred by bacteriophage. At the same time, the microbial community in the activated sludge remains exposed to the antibiotics residues, and can develop resistance against these antibiotics. A study found that tetracycline resistance genes were present in the activated sludge sampled from 15 WWTPs at different geographical locations. In particular, tetG genes were present in the highest concentration of 1.75 × 10–2 ± 2.43 × 10−3 copies per copy of 16S rRNA genes compared to the other monitored tetracycline resistance genes (e.g., tetA, tetB, Antibiotics 2013, 2 372 tetC, tetD, tetE, tetG, tetK, tetL, tetM, tetO, tetP, tetS, tetX). This high abundance of tetG genes is followed by tetC, tetA and tetS genes [28], further suggesting that the biological unit is a hot spot for antibiotic resistance genes. Zhang and coworkers demonstrated that the prevalence of antibiotic resistance in Acinetobacter isolates increased along the wastewater treatment flow path. Specifically, the number of Acinetobacter isolates that are resistant to three or more antibiotics significantly increased from 33%–72.4% from the influent to effluent samples. However, upon discharge into the river environment, the prevalence of resistant Acinetobacter isolates decreased further downstream of the discharge point (56.5%) [29], suggesting a natural die-off or loss of resistance among these isolates when exposed to the indigenous microbial community in the environment. In an earlier study that was performed to isolate Acinetobacter spp., it was noted that the prevalence of antimicrobial resistance among the 442 isolates, showed no significant difference in antimicrobial resistance for most of the antibiotics tested, except for cefotaxime and nalidixic acid. The prevalence of Acinetobacter spp. that was resistant to cefotaxime decreased from 16.9%–5% in the treated effluent than in raw influent, while the prevalence of isolates that was resistant to nalidixic acid increased from 1.5%–10% [30]. This observation suggests a loss or gain of certain mobile traits that are associated with the different antibiotic resistance genes. Alternatively, the presence of antibiotics in the activated sludge process provides a selective pressure and favorable conditions for horizontal gene transfer of antibiotic resistance genes. Besides Acinetobacter spp., viable enterococci and Enterobacteriaceae were isolated from municipal wastewater at different stages of the treatment process, and showed that many of the enterococci [9] and Enterobacteriaceae [31] were resistant to more than one antibiotic. The collective proportion of Escherichia, Shigella and Klebsiella spp. that were resistant to more than two antibiotics increased from an average 11.1% in the raw wastewater to 21.4% in the treated wastewater [31]. Similarly, the collective proportion of these Enterobacteriaceae which were resistant to three antibiotics increased from 5.5%–14.1% in the treated wastewater [31]. This observation further suggested that the conventional municipal wastewater treatment scheme does not effectively remove viable Enterobacteriaceae that are resistant to antibiotics. To counter this problem, most wastewater treatment plants opt for chlorination of effluent in an attempt to disinfect any potential viable microorganisms that may be present. The typical chlorine dosage required to achieve total coliform disinfection based on a 60 min contact time ranged from around 2.5–20 mg/L to meet a total coliform concentration of 23–200 MPN/100 mL [13]. Although chlorination is able to achieve approximately 4–6 log removal or destruction of total coliforms [13], studies have shown that a significant portion of the antibiotic resistant fecal-associated bacteria remain viable. Huang et al. [32] examined different doses of chlorination, and the subsequent impact on the inactivation and regrowth of different types of antibiotic resistant bacteria. The study concluded that high dosages of chlorination can especially enhance the recovery of chloramphenicol-resistant bacteria, while low dosages of chlorination tend to result in an increased regrowth of chloramphenicol, ampicillin and penicillin-resistant bacteria. Although it remains extremely useful to examine selective bacterial populations (e.g., Acinetobacter and Enterobacteriaceae) and their resistance traits, most of these approaches are culture-dependent and may not provide a comprehensive depiction of the actual extent of resistance among the total microbial community. This is particularly significant considering that a majority of microorganisms (i.e., up to Antibiotics 2013, 2 373 99.9%) in the environment cannot be cultivated [33–35]. For this, molecular-based approaches, particularly quantitative PCR, would be required to determine the abundances of antibiotic resistance genes that are removed by the wastewater treatment facilities, and to estimate the extent of resistance genes that are disseminated into the environment from the municipal wastewater effluent. To examine antibiotic resistance genes, the commonly used molecular methods are heavily dependent on the design of appropriate primers that can be used to detect and quantify the abundances of these genes. There are challenges involved in using this approach, primarily when examining the beta-lactamase genes. This is due to the wide diversity of beta lactamase genes that have been identified thus far [36,37]. For example, Colomer-Lluch observed five clusters of bla genes which did not CTX-M share enough conserved sequence regions to allow the design of a primer that would target all bla CTX-M genes and associated variants [27]. This means that numerous primers would have to be designed to ensure a comprehensive coverage of different resistance genes. Further search on published literature revealed that more studies are needed to collate the abundances of various types of antibiotics resistance genes present in the municipal wastewater influent and effluent (Table 1). Table 1 summarizes the quantitative measurements of antibiotic resistance genes (ARGs) that are present in raw and treated municipal wastewater. Currently, there are three ways to report the abundance of resistance genes. First, one can report the abundance of resistance genes by normalizing against the volume of sample that was processed and extracted for its genomic DNA. Second, abundance of resistance genes are normalized against the total copy of 16S rRNA genes that were detected in the sample. Third, one can report the abundance of resistance genes that was normalized against the mass of genomic DNA utilized for quantitative PCR. All three methods would give slightly different values and hence interpretations [38]. For example, the first method would be useful when performing quantitative microbial risk assessment as it provides an estimate of the abundance of resistance genes at which a subject is exposed to when a known amount of treated wastewater was reused. The second and third method provided an estimate of the ratio of resistance genes to the total biomass and bacteria, respectively, which are present in the sample. Normalization to 16S rRNA genes and biomass would account for minor variations in sample processing, such as differences in DNA extraction efficiency, and in turn allow appropriate comparisons to be made among different samples. Furthermore, an increase in the abundance of resistance genes against the 16S rRNA genes would indicate that while the wastewater treatment process is efficient in removing most bacteria, antibiotic resistance genes, possibly along with their resistant hosts, were not effectively reduced by the same treatment schematics. To illustrate, a conventional activated sludge process achieved less than 1 log reduction in the ratio of tetA and tetC genes against the amount of genomic DNA [26]. Similarly, different resistance genes have different removal efficiency by the same treatment schematic. Auerbach and coworkers reported a 3 log removal of tetQ per ng DNA while tetG has only 2 log removal in the same Wisconsin wastewater treatment plant [39]. Antibiotics 2013, 2 374 Table 1. Abundance of antibiotic resistance genes that was present in untreated municipal wastewater and in treated effluent. Abundances were determined by quantitative PCR. Abundance in final Gene Abundance in raw Treatment Geographical Antibiotics discharge/impacted Ref. class water procedure location environment Beta-lactam 106.15/mL sample 105.61/mL sample Massachusetts, bla 102.51/ng DNA 102.38/ng DNA AS + Cl [38] TEM-uni USA 10−4.74/16S 10−3.64/16S 105.35/mL sample 103.45/mL sample South Carolina, bla AS + Cl [40] M-1 101.10/ng DNA 102.03/ng DNA USA bla 10−1.22–101.26/ng DNA ND–102.2/ng DNA WWTP, not vim Germany [41] ampC 100.34–102.66/ng DNA 10−0.27–101.20/ng DNA specified NA ND–101.84/mL sample AS + P and N Sabadell, Spain NA ND/mL sample Galatone WWTP Nardò, Italy ampC Torreele, NA ND/mL sample AS + UF + RO Belgium NA ND–102.06/mL sample AS + P and N Sabadell, Spain NA ND/mL sample Galatone WWTP Nardò, Italy bla [42] shv-5 Torreele, NA ND/mL sample AS + UF + RO Belgium NA ND–102.75/mL sample AS + P and N Sabadell, Spain NA ND/mL sample Galatone WWTP Nardò, Italy mecA Torreele, NA ND/mL sample AS + UF + RO Belgium bla 104.4/mL sample 103.8/mL sample TEM Barcelona, bla 103.3/mL sample 102.1/mL sample Not specified [27] CTX-M Spain mecA 103.7/mL sample 102.2/mL sample 101.7/mL sample 100.7/mL sample Gothenburg, mecA AS + TF [43] 10−0.07/ng DNA 10−0.59/ng DNA Sweden Macrolide NA ND–104.42/mL sample AS + P and N Sabadell, Spain ermB NA ND–103.13/mL sample Galatone WWTP Nardò, Italy [42] Torreele, NA ND–103.28/mL sample AS + UF + RO Belgium Water solids, ermB ~109.70/mL sample ~107.70/mL sample aerobic digestor Minnesota, [44] ~10−2.30/16S ~10−3.60/16S for approximately USA 3 months ~102.3–104.48/mL ~107.48/mL sample ermB sample ~10−3.52/16S Secondary ~ND–10−3.82/16S effluent, activated Shafdan, Israel [45] ~103.48–105.3/mL ~107.78/mL sample sludge ermF sample ~10−3.22/16S ~10−3.05–10−2.52/16S Antibiotics 2013, 2 375 Table 1. Cont. Abundance in final Gene Abundance in raw Treatment Geographical Antibiotics discharge/impacted Ref. class water procedure location environment Tetracycline 105.37−107.4/mL sample ND–103.63/mL sample tetW ~10−3.12/16S ND/16S AS/OD/RBCs/MB Michigan, USA [46] 105.51–107.61/mL sample ND–103.96/mL sample R + UV/Cl tetO ~10−3.12/16S ND/16S 107.2–109/mL sample 103.9–106.2/mL sample tetQ 105.3–106.8/ng DNA 103.7–105.4/ng DNA AS/P and Wisconsin, [39] 106.4–107.8/mL sample 104.2–105.9/mL sample N/UV/Cl USA tetG 104.5–105.7/ng DNA 104–105/ng DNA 108.13–108.3/mL sample ND–104.12/mL sample tetC 105.45–105.65/ng DNA ND–103.57/ng DNA Hong Kong, AS + Cl [47] 107.78–108.2/mL sample ND-104.33/mL sample China tetA 105.09–105.57/ng DNA ND–103.78/ng DNA 107.7/mL sample 106.15/mL sample tetA 105.55/ng DNA 105.24/ng DNA AS Nanjing, China [26] 107.91/mL sample 106.14/mL sample tetC 105.76/ng DNA 105.23/ng DNA NA ND–104.44/mL sample AS + P and N Sabadell, Spain 102.93–104.58/mL NA Galatone WWTP Nardò, Italy tetO sample [42] Torreele, NA ND–105.02/mL sample AS + UF + RO Belgium ~108.85/mL sample ~107.85/mL sample tetA ~10−3.15/16S ~10−3.46/16S Water solids, ~109.78/mL sample ~107.95/mL sample aerobic digestor Minnesota, tetW [44] ~10−2.22/16S ~10−3.35/16S for approximately USA ~108.7/mL sample ~109.48/mL sample 3 months tetX ~10−3.3/16S ~10−1.82/16S ~103, 101.7, 102/mL sample Secondary tetO NA ~10−2.3, 10−2.6, effluent, 10−2.7/16S chlorinated Western USA [48] ~103.6, 102.3,102/mL effluent, sample dechlorinated tetW NA ~10−1.7, 10−2, effluent 10−2.7/16S Secondary ~107.3/mL sample ~ND−103/mL sample tetO effluent, activated Shafdan, Israel [45] ~10−3.7/16S ~ND–10−4.4/16S sludge Antibiotics 2013, 2 376 Table 1. Cont. Abundance in final Gene Abundance in raw Treatment Geographical Antibiotics discharge/impacted Ref. class water procedure location environment tetM 10−3.87–10−2.42/16S Rural domestic tetO 10−4.41–10−3.24/16S sewage treatment NA tetQ 10−4.64–10−2.8/16S system, usually tetW 10−3.16–10−2.03/16S anaerobic digestor Hangzhou, [49] tetM ~10−2.70–10−2.30/16S Urban WWTP, China tetO ~10−3.00–10−2.70/16S usually oxidation NA tetQ ~10−2.82–10−2.00/16S ditch or anaerobic tetW ~10−1.70–10−1.30/16S oxic zones Sulfonamide 104.37–106.75/mL 105.46–107.54/mL sample AS/OD/RBCs/MB sul-I sample Michigan, USA [46] ~10−3.4/16S R + UV/Cl ~10−3.4/16S ~106.4/mL sample ~106.5/mL sample sul-I ~10−1.52/16S ~10−1.1/16S Lausanne, Not specified [50] ~105.6/mL sample ~105.5/mL sample Switzerland sul-II ~10−2.3/16S ~10−2/16S Water solids, ~108.90/mL sample ~108/mL sample aerobic digestor Minnesota, sul-I [44] ~10−3.10/16S ~10−3.30/16S for approximately USA 3 months ~104.9, 103.7, 103.9/mL sample Secondary sul-I NA ~10−0.4, 10−0.6, effluent, 10−0.8/16S chlorinated Western USA [48] ~104.6, 102, 101.9/mL effluent, sample dechlorinated sul-II NA ~10−0.7, 10−2.3, effluent 10−2.8/16S ~104.78–105.48/mL ~108.48/mL sample sul-I sample ~10−2.4/16S Secondary ~10−2.52–10−1.7/16S effluent, activated Shafdan, Israel [45] ~103.48–104.88/mL ~108.30/mL sample sludge sul-II sample ~10−2.7/16S ~10−3.3–10−2.4/16S sul-I ~10−2.70–10−1.70/16S Rural domestic sewage treatment NA sul-II ~10−2.52–10−1.15/16S system, usually anaerobic digestor Hangzhou, [49] sul-I ~10−2.15–10−1.7/16S Urban WWTP, China usually oxidation NA sul-II ~10−2.00–10−1.70/16S ditch or anaerobic oxic zones