Since ancient times, humans have randomly disposed of waste into the environment, such as in rivers and cesspits. The industrial revolution of the late eighteenth and early ninteenth centuries was a period that saw increased disposal of toxic organic chemicals by direct release into the environment. Many of these toxic molecules had antimicrobial activity, and it can be assumed that microbes resistant to these toxins multiplied in such environments. As a modern example, one can cite the concentrations of heavy oils that were dumped near detection stations in the distant early warning line at the end of the Second World War. These sites are now excellent sources of bacteria with enhanced biodegradation capacities and have been extensively studied in recent years.
Following the discovery of the chemically synthesized sulphonamides and trimethoprim and the identification of dual resistance in 1969, the subsequent and most disastrous environmental pollution has come from the disposal of antibiotic production wastes in various forms. These discarded products were developed as food supplements for farm animals to promote weight gain in all aspects of animal and fish husbandry worldwide. The amounts of antibiotics and antibiotic wastes disposed in this way cannot be accurately identified. However, according to recent estimates by the Union of Concerned Scientists in the United States, antibiotic use for nontherapeutic purposes in three major livestock sectors (chickens, cattle, and swine) was about eight times more than the consumption for human medicine (Mellon et al., 2001).
In the past 50 years, we have seen the rapid evolution of a new plague—that of worldwide antibiotic resistance. Though not a disease in itself, antimicrobial resistance (AR) results in the failure to effectively prevent and treat many diseases, leading to widespread untreatable microbial infections and greatly increased morbidity and mortality: a plague of resistance genes (Davies and Davies, 2010). The global use of antibiotics at low cost, auto medication, and short duration of treatment has accelerated, extended, and expanded the spectra of resistance worldwide. The earth has been continuously bathed in a dilute solution of antibiotics for more than half a century.
Aquatic ecosystems have been identified as hotspots of resistance mechanisms (Rizzo et al., 2013). This is due to the large diversity of pathogenic and commensal microorganisms and the continuous discharge of antibiotic resistant bacteria (ARB) and antibiotic resistance genes (ARGs) into these environments. As part of aquatic ecosystems, urban wastewater treatment systems (collecting sanitary sewage, hospital effluents, and storm water runoff) possess all the components required to ensure the acquisition of all varieties of resistance genes. The antimicrobials present in wastewater due to incomplete degradation by humans and animals, disposal of unused drugs, and runoff losses from land application, together with environmental and pathogenic bacteria in nutrient‐rich engineered systems, provide all the necessary requirements to support a breeding ground for horizontal gene transfer and the propagation of resistance genes (Davies and Davies, 2010; Ferreira da Silva et al., 2006; Kim and Aga, 2007; Lefkowitz and Duran, 2009).
Since 1890 with the building of the first biological wastewater treatment plant (WWTP) in Worcester, Massachusetts, advances in wastewater treatment technology have been improving the efficient removal of biodegradable organic pollutants. Currently, enhanced biological phosphorus removal processes have not only enabled the removal of traditional carbonaceous contaminants but also reduced phosphorus concentrations to very low levels (<0.1 mg/L) in the effluent discharge (Zuthi et al., 2013).
Over the past 15 years, increasing attention has shifted toward the identification and removal mechanisms of micropollutants from wastewater and sludge. Micropollutants are persistent organic or mineral substances such as pharmaceuticals and personal care products, detergents, and pesticides whose discharge, even at very low concentrations, is a constant growing environmental contamination (Luo et al., 2014).
Despite the evolution of wastewater treatment technologies from conventional to advanced treatment configurations, existing urban biological wastewater treatment systems are not designed to remove micropollutants and ARGs. Studies on antibiotics as emerging classes of micropollutants have confirmed the high frequency of antimicrobial resistant genotypes as well as ARB in wastewater treatment systems, including constructed wetlands and WWTPs (Martins da Costa et al., 2006; Kim et al., 2010; Volkmann et al., 2004; Luczkiewicz et al., 2010; Reinthaler et al., 2003).
In a landmark series of papers published between 2003 and 2009, Szczepanowski and colleagues presented the first extensive DNA sequence–based screening of a large set of known ARGs in samples of activated sludge and the final effluent of a WWTP in Bielefeld‐Heepen, Germany. This comprehensive survey identified 140 different clinically relevant antimicrobial resistant genotypes and contaminants. From these investigations, it is evident that such treatment systems may play important roles in the development and assortment of multidrug‐resistant (MDR) bacteria among complex populations.
The occurrence of ARB and ARGs in the two main by‐products of wastewater treatment systems (biosolids and effluent discharge) has been reported frequently. Currently, effluent water quality standards, prior to discharge, are limited to controlling the concentrations of carbonaceous biochemical oxygen‐demanding matter, suspended solids, total residual chlorine and un‐ionized ammonia. There exist no regulatory guidelines to monitor and control the levels of ARGs in bacteria and extracellular DNA from lysed microbial cells in the effluent discharge. Accordingly, studies have reported that antibiotic resistance determinants and MDR pathogens are transported from the effluent to the receiving water (Iwane et al., 2001; Galvin et al., 2010; Goñi‐Urriza et al., 2000). For example, LaPara et al. (2011) showed that the quantities of three tetracycline resistance genes were significantly higher in a tertiary treated effluent discharge than in receiving water samples in the St. Louis River, Duluth‐Superior Harbor, and Lake Superior, USA.
Despite the evidence for the occurrence of resistance genes in effluent discharge points, the overall impact of treated wastewater applications on irrigation processes is unclear. Some studies have observed an increase in soil microbial activity and biomass after irrigation by treated wastewater as shown by a shift in the composition of soil bacterial communities (Oved et al., 2001; Broszat et al., 2014). However, recent studies have observed no significant impact on AR in the wastewater‐irrigated soil microbiome (Gatica and Cytryn, 2013; Negreanu et al., 2012).
The presence of ARB and ARGs in biosolids‐amended soils is well documented (Brooks et al., 2006; Rahube et al., 2014). Biosolids are the treated and stabilized nutrient‐rich organic residuals produced as a by‐product of wastewater treatment and widely used as fertilizer to stimulate plant growth (Lu and Stoffella, 2012). Recent studies have demonstrated that complementary technologies such as aerobic digestion and lime stabilization can be used as approaches to reduce the quantities of ARGs in biosolids (Munir et al., 2011). However, ARG concentrations and corresponding decay rates can be variable depending on the application methods, biosolids treatment reactor design, storage conditions, the specific ARGs involved, and the frequency of biosolids application (Burch et al., 2013; Miller et al., 2014).
Although ARB and genes encoding antibiotic resistance have been commonly detected in wastewater and the by‐products of treatment systems, the role of wastewater treatment processes in the dissemination of AR is not clear. In recent years, a number of studies have...
