The Emergence of Antibiotics Resistance Genes, Bacteria, and Micropollutants in Grey Wastewater

Abstract

The reuse of household greywater is increasing globally. Wastewater and greywater treatment processes are not fully effective in removing all contaminants, such as emerging micropollutants, antimicrobial-resistant bacteria, and antibiotic resistance genes. The dynamics between emerging micropollutants and antibiotic resistance genes in greywater treatment systems are complex. Thus, this review aims to analyze the current knowledge on sources, spread, and the fate of emerging micropollutants, antibiotic-resistance genes, and antimicrobial-resistant bacteria in microbial communities of greywater and downstream recipients. The fate of antimicrobial resistance and emerging micropollutants from greywater in the environment has not been determined. More studies are needed to identify the mechanism/s involved in the degradation of emerging micropollutants and the presence of transformation pathways in the microbial metagenome. In the review, we aim to describe the link between the persistence of emerging micropollutants and the emergence of antimicrobial resistance. We showed that the effect of irrigation with treated wastewater was variable. In addition, we tried to summarize the impact of emerging micropollutants on bacteria and their fate in the soil microbiome, demonstrating that emerging micropollutants induce changes in the diversity of soil bacteria. The fate and transport of emerging micropollutants, antimicrobial-resistant bacteria, and antibiotic resistance genes can vary with soil properties. It is, therefore, necessary to better understand how widely antibiotic-resistance genes are disseminated.

1. Introduction

Water is a finite resource with a growing demand. Many water sources are depleting, becoming polluted, or both. Globally, 2.3 billion people live in countries with water stress, of which 733 million live in high and critically water-stressed countries [1]. These people live without safe drinking water and sanitation, even though both services have long been defined as human rights. This water scarcity has led to an increased interest in wastewater reuse, comprising a reduction of domestic water consumption and decreasing the volume of wastewater that has to be treated [2]. Irrigation with greywater (GW) is one of the water reuse methods which is currently widely practiced [3]. This review defines greywater as all non-toilet wastewater resulting from households. As such, it contains a lower organic load, and the microbial quality is expected to be higher than in blackwater; therefore, it can be treated relatively simply for reuse, even locally [4]. Recycling GW can reduce potable water demand by up to 50% [5,6]. Nevertheless, it is established that raw greywater is contaminated with pathogens, synthetic chemicals, and antimicrobial compounds and should, thus, be treated before reuse [7].

The term “antibiotic” has been used to characterize molecules that inhibit or kill microorganisms, classified based on their mechanisms of action [8]. Residues of antibiotics in wastewater and greywater can be divided into nine commonly used classes; (1) β-lactams (e.g., penicillin, cephalosporin, and amoxicillin), (2) Fluoroquinolones (e.g., ciprofloxacin, norfloxacin, and ofloxacin) (3) Lincosamides (e.g., pirlimycin and clindamycin) (4) Macrolides: erythromycin, azithromycin, clarithromycin, and tylosin) (5) Sulfonamides (e.g., sulfamethazine and sulfamethoxazole) (6) Reductase inhibitor (e.g., trimethoprim) (7) Tetracycline family (e.g., tetracycline, minocycline, and oxytetracycline) (8) Glycopeptide (e.g., vancomycin), and (9) Amphenicol (e.g., chloramphenicol) [9]. Residues of antibiotic compounds can impose pressure on the wastewater microbiome, even in concentrations below minimal inhibitory concentrations (MIC) [10]. In other words, the antibiotics can increase resistance even below concentrations needed to prevent the visible growth of the microorganism.

1.1. Development of Resistance

Resistance is the ability of a microbe to become insensitive to an antibiotic compound [11]. Microbes have evolved several protective (resistant) mechanisms to overcome the effectiveness of antibiotic compounds. It includes the degradation of the antibiotics (β-lactamases), alteration of the antibiotic target structure (DNA-gyrase), inactivation (by an aminoglycoside N-acetyltransferase of fluoroquinolone), increased efflux pumps activity (export of a drug out of the microorganism) and protection of the target (by DNA-binding proteins, for example, known as Qnr). Most antibiotic-resistant bacteria are Gram-negative bacteria [12], mainly because of their relatively impermeable cell wall [13]. The cell wall is surrounded by an external lipid membrane that restricts the diffusion of hydrophobic compounds [14]. Furthermore, by changing their cell wall characteristics, such as mutations in porins, Gram-negative bacteria can obtain resistance [12].

Bacteria can possess intrinsic resistance to antibiotics or may acquire resistance by acquiring new genetic materials from resistant strains through horizontal gene transfer (HGT) mechanisms [15]. HGT includes transformation (uptake and recombination of naked DNA), transduction (phage mediated), and conjugation (commonly through mobile genetic elements like plasmids (MGEs) [16]. Mutation and selection, in combination with gene transfer mechanisms, allow the bacteria a faster adaptation to antibiotics stress and enable the transfer of resistance among different species. Adding new genes into bacterial chromosomes through integrons is a common mechanism for acquiring resistance [17]. All the genes responsible for antibiotic resistance are termed “resistome” [18]. It was acknowledged long ago that other mechanisms could help bacteria survive exposure to antibiotic, such as tolerance [19]. However, this review will not focus on additional pathways in which bacteria obtain resistance to antibiotics.

1.2. Prevalence of Antibiotic-Resistant Bacteria and Antibiotic-Resistance Genes in GW

Hendriksen et al. [20] analyzed 79 samples of untreated urban sewage in 60 countries, concluding that the dominant bacterial genera were typically fecal, including Escherichia, Bacteroides, Bifidobacterium, Streptococcus, and Faecalibacterium. Pathogens including Klebsiella pneumoniae, Pseudomonas aeruginosa, Salmonella spp., Enterovirus, Escherichia coli, Giardia, Staphylococcus aureus, Clostridia, and Legionella sp. were detected in untreated domestic greywater in England, Australia, U. S., Hungary, Israel, Uganda, and France [21,22,23,24]. In addition, many studies have revealed pathogenic bacteria in treated domestic wastewater worldwide, such as Salmonella enterica, Listeria monocytogenes [25], E. Coli [26], Aeromonas hydrophilia, and Aeromonas veronii [27].

Great interest has been explicitly aroused about antibiotic-resistant bacteria (ARB) and antibiotic-resistance genes (ARGs) in greywater [28,29,30]. Many studies monitored the occurrence of ARGs and ARBs in wastewater [27,31,32,33]. For example, Szczepanowski et al. [32] report the occurrence of resistance genes to macrolide, aminoglycoside, b-lactam, chloramphenicol, tetracycline, fluoroquinolone, sulfonamide, trimethoprim, and rifampicin, including multidrug resistance (MDR) genes. The article summarizes 140 clinically relevant ARGs in wastewater treatment plants. Randall et al. [34] report ARGs in strains of Salmonella enterica, including genes such as aadA1, aadA2, aadB, aphAI-IAB, bla(Tem), bla(Carb2), cat1, cat2, dhfr1, floR, strA, sul1, sul2, tetA(B) tetA(A), and tetA(G). Another example is the research of Hayatgheib et al. [35], detecting 44 ARGs in Aeromonas strains. Thus far, 211 isolates were resistant to colistin, quinolones, florfenicol, oxytetracycline, and trimethoprim-sulfamethoxazole. Alongside that, several studies have reported the occurrence of ARGs in small-scale GW reuse systems, including tet39, blaCTX-M-32, qnrS, and sul1 resistance genes [23,36,37]. Troiano et al. [36] identified Pseudomonas putida in treated greywater, which is clinically resistant to cephems and carbapenems. Yomoda et al. [38] reported that resistance genes of P. putida were transferable by transformation or conjugation to recipient strains of Pseudomonas aeruginosa. Overall, it is recognized that wastewater treatment facilities are potential hotspots for the spread of ARGs [39].

In wastewater treatment, the risk of conventional pathogens and common pollutants from the reuse of GW has been widely reviewed. The prevalence of ARBs in greywater, mitigation of ARGs, and potential risk to humans and the environment by greywater-reuse practices should be given more attention. Moreover, understanding the possible effect of emerging micropollutants (EMPs) on the multidrug resistance of bacteria and cross-resistance, and the dissemination of antimicrobial chemicals in the environment is crucial. In addition, we tried to summarize the effect of emerging micropollutants on bacteria and their fate in the soil microbiome. Therefore, this review surveys research focused on the most occurring EMPs in GW, and the risk of greywater reuse practices on the environment.

2. Emerging Pollutants

Maintaining good personal hygiene is essential to prevent infectious diseases from occurring and spreading. Everyday practices such as cleaning objects used often, washing face and hair, and brushing teeth, require using appropriate synthetic chemicals (personal health care products, PCP). EMPs are compounds that have recently been classified as harmful to the environment and, consequently, the health of human beings. One specific group is micro-pollutants (M.P.s): contaminants found in trace concentrations (microgram to nanogram per litter or kg). The most detected micropollutants in greywater are Triclosan (biocide), Methylparaben, and Propylparaben (preservatives), Galaxolide and Tonalide (fragrances), as well as Oxybenzone and Octocrylene (U.V. filters) and Benzalkonium chloride. Biocides are active chemicals that control the growth of bacteria or kill them [40]. Preservatives are compounds that inhibit the growth of any infectious microorganisms that may be present. Fragrance ingredients are extensively used in PCPs. Ultraviolet (UV) filters are compounds that block or absorb ultraviolet light [41]. Benzalkonium chloride is used primarily as a disinfectant and is a common ingredient in domestic applications like personal hygiene products or fabric softeners [42].

Non-antibiotic household chemicals are used more extensively than antibiotics on a global scale. Recent research identified the occurrence of bacteria and EMPs in household greywater as a potential aspect responsible for the accumulation of ARB [43,44]. Synthetic chemicals can act in an unspecific manner, targeting different sites or processes in bacterial cells and causing harm to multiple bacteria cells [45,46]. This mode of action is different from antibiotics, which attack specific targets. Nevertheless, bacterial resistance to synthetic chemicals can be created by the exact protective mechanisms against antibiotics.

Table 1. Synthetic chemicals with their basic physical and chemical properties.

Compound	Molecular Formula	Chemical Structure	Molecular Weight (g/mol)
Tonalide	C18H26O		258.4
Galaxolide	C18H26O		258.4
Oxybenzone	C14H12O3		228.2
Octocrylene	C24H27NO2		361.5
Methylparaben	C8H8O3		152.2
Propylparaben	C10H12O3		180.2
Triclosan	C12H7Cl3O2		289.5
Benzalkonium chloride	C22H42CINO		372.0

presents the synthetic chemicals with their basic physical and chemical properties. These non-antibiotic household chemicals could induce different bacterial responses, including the regulation of gene expression, stress, ROS formation, and SOS response. These responses either promote HGT or increase gene mutation frequency [47].

2.1. Cross-Resistance

Concerns about the effect of antimicrobials on the increasing resistance to antibiotics of bacteria were raised more than 50 years ago [48]. Giuliano and Rybak [44] showed that there is a potential link between triclosan and antibiotic resistance. Lu et al. [49] report bacterial mutants resistant to quinolone and mupirocin that have decreased susceptibility to triclosan. Exposure to benzalkonium chloride and triclosan resulted in increased resistance to erythromycin and ciprofloxacin strains of Campylobacter jejuni and Campylobacter coli [50]. Triclosan exposure in strains of E. coli and P. aeruginosa increased ten times the resistance to chloramphenicol and tetracycline [51]. However, there is still a lack of evidence on the role of EMPs and their direct effect on antibiotic resistance.

2.2. Ecotoxicological Effects of EMPs

Triclosan (TCS) is a common antimicrobial chemical in numerous PCPs (soaps, sanitizers, and toothpaste) [52] and is widely detected in aquatic environments at μg/L [53,54] to mg/L level [55]. It can serve as an external pressure to co-select for triclosan resistance and antibiotic resistance in many bacteria [56,57,58]. Triclosan induces oxidative stress, causing genetic mutations in a few genes, such as marR, frdD, fabI, acrR, and soxR [59,60,61]. The gene fabI is an acyl carrier protein reductase gene, a key enzyme in fatty acid synthesis in bacteria [62]. The interference with fatty acid synthesis results in modifications of the membrane structure, which causes less antibiotic uptake [63]. Subinhibitory concentrations of triclosan decrease the susceptibility of E. coli to ciprofloxacin, kanamycin, and gentamicin due to alteration of the membrane structure and biofilm formation [64]. Likewise, Stenotrophomonas maltophilia exposed to triclosan resulted in overexpression of the multidrug efflux pump SmeDEF and reportedly reduced the susceptibility to chloramphenicol, tetracycline, and ciprofloxacin [65]. Similarly, Salmonella enterica exposed to triclosan with increasing concentrations showed overexpression of the AcrAB efflux pump and reportedly reduced susceptibility to chloramphenicol, tetracycline, and ampicillin [66]. An efflux pump overexpression suggests the co-selective potential for more antimicrobial chemicals. In addition, E. coli exposed to triclosan also resulted in overexpression of the multidrug efflux pump and a transcription of genes encoding beta-lactamases. In contrast, the expression of genes related to membrane permeability decreases [61]. Hartmann et al. [67] documented that exposure to antimicrobials methyl-, ethyl-, propyl-, butylparaben, triclocarban, and triclosan increases ARGs in the microbiome.

Parabens are widely used as preservatives in many pharmaceutical, food, and cosmetic products due to their low toxicity [68]. They are often used in small amounts and primarily prevent bacteria growth and prolong shelf life. The mode of action against microorganisms mostly interferes with cellular membrane transfer processes. The effectivity of the paraben is correlated with the size of the chemical; propylparaben is considered more active against most bacteria than methylparaben [69]. Propylparaben specifically induces the permeabilization of bacterial membranes, causing the release of potassium [70,71]. The mechanisms of microbial resistance to parabens need to be better understood. Parabens are less active toward Gram-negative than Gram-positive bacteria [72,73]. Paraben resistance has been linked to the cell wall characteristics and non-specific efflux systems [74]. However, so far, only a few cases of resistance to parabens have been reported, occurring in the specific strains of P. aeruginosa, Burkholderia cepacia, and Cladosporium resinae [68]. Wu et al. [75] reported biodegradation of methyl- and propylparaben under aerobic and anoxic conditions, and benzoic acid was identified as one of the significant degradation products, thus reducing the efficacy of the compound.

Tonalide and galaxolide are the most used synthetic fragrance compounds in various PCPs, such as detergents, perfumes, deodorants, skin creams, and soaps [76]. A tonalide concentration has been detected in wastewater treatment plants (WWTP), ranging between 0.086 and 12.5 μg/L and 0.043 and 16.6 μg/L for galaxolide [77]. They are moderately soluble in water and thus increase the possibility of accumulation in the environment [78]. Tonalide and galaxolide are volatile lipophilic compounds and can, therefore, relatively easily penetrate through the cell wall of microorganisms. They are more active against Gram-positive than Gram-negative bacteria [79]. Tonalide and galaxolide disrupt and damage the structure of the membrane, resulting in a loss of ions, collapse of the proton pump and cytoplasm leakage, enzyme inhibition, and proton exchange disruption [80]. The metabolic pathway of hydroxylation is mainly causing the biotransformation of galaxolide, which is attributed to the existence of the cytochrome P-450 enzymes, which are linked to the inactivation of antibiotics [81].

Oxybenzone and octocrylene are filters that absorb U.V. radiation between 280 and 400 nm. U.V. filters are the main components of sunscreen due to their absorbing properties, but they are also found in other industrial products such as plastics and paints. Lozano et al. [82] were the first to analyze the effect of these compounds on microbes, reporting that they affected only gram-negative microbes. A correlation has been detected between genome size and the appearance of resistance mechanisms [83].

Quaternary ammonium compounds (QACs) are a group of chemicals found in most household cleaning products because of their different functions. They can act as disinfectants, surfactants, or preservatives [84]. Several genes, such as qacE, qacE11, qacF, qacG, and qacH, have been reported to confer resistance to QACs in Gram-negative bacteria, with qacE11 being the most widespread [85]. These genes belong to the small multidrug resistance (SMR) family [86], and their resistance to QACs is efflux-mediated [87]. The use of QACs drives the spread of class I integrons, responsible for a significant part of antimicrobial resistance in Gram-negative bacteria [88,89]. Benzalkonium chlorides (BAC) are the most commonly used QACs [90]. Isolates of P. aeruginosa exposed to increasing concentrations of BAC caused mutations in the pmrB gene and physiological adaptations that contributed to a higher tolerance to antibiotics [91]. Additionally, Guerin et al. [92] report the susceptibility of Listeria monocytogene to various antibiotics such as ciprofloxacin, gentamicin, or kanamycin after exposure to BAC. Efflux pump expression most likely causes antibiotic resistance to BAC, accompanied by the minor role of reduced membrane permeability [93]. Efflux genes such as qacG, acrA, qacH, and acrB have been identified in bacteria resistant to BAC [94].

Table 2. EMPs and their effect on bacteria and the defense mechanism of the bacteria.

Active Compound	Disinfectant Working Mechanism	Bacterial Adaptation to Disinfectant	References
Triclosan	- Oxidative stress in bacterial cell - Genetic mutation of the enoyl-acyl carrier protein (ACP) reductase genes - Fatty acid synthesis disruption	- Active efflux pumps - Membrane permeability decrease - Biotransformation, horizontal gene transfer - Increased target expression (overexpressed genes mufA1 and mufM)	[61,95,96,97,98]
Methyl- and propylparaben	- Membrane disruption - Cell leakage - Induction of potassium efflux	- Change of cell wall characteristics - active efflux pumps	[69,70,71]
Tonalide and Galaxolide	- Membrane disruption - Enzyme inhibition - Proton exchange disruption.	- Existence of cytochrome P-450 (biotransformation)	[80,81]
Oxybenzone and octocrylene	- General toxic effects like reduced growth, energy, and DNA metabolism.	- Multidrug transporters - ROS responsive elements - Periplasmic stress response regulons	[82,83,99]
Benzalkonium chlorides	- Spread of intI1 gene - Cytoplasmic membrane damage	- Increasing horizontal gene transfer - Downregulation of membrane porins - Overexpression of efflux pumps	[100,101,102,103]

summarizes the EMPs their effect on bacteria and the defense mechanism against EMPs. In summary, non-antibiotic chemicals induce antibiotic resistance, and the cell wall is the first encounter between the bacteria and the chemical and is, thus, an essential mechanism of resistance.

3. Irrigation with Treated GW

Existing GW treatment systems are not designed to remove EMPs, ARBs, and ARGs, which remain in the effluent. According to Troiano et al. [36], the filter bed biofilm contributes to the ARB community in the treated effluents. Thus, the reuse of GW may pose a severe public health risk by disseminating resistance, especially if humans are exposed to places where ARBs are present [104]. ARBs may persist in the environment and disperse ARGs to other bacteria [33]. Evidence shows that the environment plays a vital role in AMR development, dissemination, and transmission [105]. Only a handful of research has focused on the effect of irrigation with treated GW or treated wastewater (TWW) on ARG and ARB in soils. Soil is considered one of the largest ARG environmental reservoirs [106]. Treated GW used for garden irrigation and landscaping introduces significant quantities of ARGs into soil habitats that might have previously been unexposed [107,108,109]. Moreover, treated GW irrigation might promote the maintenance of ARGs in soils; however, Pepper et al. [110] stated that antibiotic resistance levels in the soil are increased temporally by anthropic activities, but their persistence is not guaranteed. It should be emphasized that even undisturbed soils contain highly abundant and diverse levels of ARB due to antibiotic-producing soil fungi or bacteria [110]. A systematic overview of studies is given in

Table 3. A summary of studies investigating the effects of treated greywater on the presence of ARGs and ARBs in the soil.

Research	Location	Findings	Reference
Field irrigated with TWW or fertilizer-amended water	Israel	- in soils irrigated with fertilizer-amended water, the Nitrosospira populations were dominant - Nitrosomonas populations were dominant in TWW irrigated soils.	[111]
Fields subjected to short wastewater irrigation regime for more than 15 years	Senegal	- no change in the total bacterial community - a significant change in the AOB community by groundwater supply.	[112]
short- and long-term irrigation with municipal wastewater	Tunisia	- TWW irrigation led to changes in the genetic structure of bacterial communities - significant increase in microbial densities with TWW irrigation	[113]
Field irrigated with TWW or potable water.	Israel	- irrigation with TWW did not result in the transfer of fecal indicator bacteria or microbial pathogens to the irrigated soil	[114]
Fields irrigated with either TWW or freshwater	Israel	- soil irrigated with freshwater contained a similar or higher relative abundance of ARGs and ARBs	[115]
Household gardens irrigated with either treated GW or freshwater	Israel	- soils irrigated with GW and freshwater had the same pathogen levels	[116]
Five fields irrigated with TWW directly or from rivers that receive effluent, compared to fields without irrigation	China	- no difference in ARB between irrigated and non-irrigated soils, while the concentration of ARGs in irrigated soil was higher	[117]
Soil from water storage basins irrigated either with TWW or groundwater	USA	- no increase in ARGs regardless of irrigation treatment - more MDR bacteria were detected in sediments from groundwater-filled ponds than in wastewater-exposed sediments	[118]
Soil columns in plastic containers irrigated with TWW and synthetic freshwater	USA	- no significant differences in microbial diversity - more abundancy of Pseudomonas, Legionella, and Acinetobacter with TWW irrigation	[119]
Household gardens irrigated with either treated GW or freshwater	Israel	- no difference in the abundance of tetracycline-resistant bacteria	[36]
The field was irrigated with 92% TWW from 10 wastewater treatment plants, compared to the field with groundwater irrigation	Spain	- increase in the relative abundance of tetM, mecA, and blaOXA-58 genes with TWW irrigation, while the genes blaTEM and sul1 were unaffected	[120]
Two experimental plots irrigated with either freshwater or TWW	Israel	- almost no correlation between ARG abundance in irrigation water and those detected in soil - intI1 may not always be a reliable indicator to determine the impact of TWW irrigation	[121]
Field irrigated with either TWW or freshwater	Germany	- increase in the relative abundance of some ARGs, including sul1 and intI1, with TWW irrigation - no difference in bacterial load	[109]
Soil microcosms with either freshwater or TWW	Israel	- in freshwater-irrigated soil, Gram-positive bacteria Bacillaceae strongly proliferated - TWW irrigation stimulated the growth of the Enterobacteriaceae and Moraxellaceae families	[122]

according to the year of publication.

Both Oved et al. [111] and Ndour et al. [112] assessed the fate of ammonia-oxidizing bacteria (AOB) in short- and long-term irrigated soils. In both studies, the results suggested that TWW irrigation produces shifts in the AOB population in the soil. Oved et al. [111] revealed that Nitrosomonas strains dominate the AOB population in TWW-irrigated soils. In contrast, the AOB populations were dominated by Nitrosospira strains in soils irrigated with freshwater. Nevertheless, no apparent changes were observed in community function despite shifts in the microbial community. Additionally, Ndour et al. [112] saw no differences in microbial activity or microbial biomass comparing the two treatments. Long-term irrigation with TWW led to changes in the genetic structure of bacterial communities, as reported by a study by Hidri et al. [113]. Orlofsky et al. [114] assessed the survival of bacterial pathogens in GW-irrigated soils. While many pathogens were present in GW, no significant differences were observed between GW-irrigated soils and freshwater-irrigated soil. Negreanu et al. [115] compared fields irrigated with treated wastewater to fields irrigated with freshwater. The relative abundance of ARBs was similar or higher in the freshwater-irrigated soils. The relative abundance of ARGs was higher in the freshwater-irrigated soils at three sites and higher in wastewater-irrigated soils at the fourth site. Benami et al. [116] investigated private gardens irrigated with GW treated by recirculating vertical flow constructed wetland. Several pathogens were detected (Klebsiella pneumoniae, Shigella spp., Salmonella enterica) in soils irrigated with the treated GW, but the same pathogen levels were also detected in soils irrigated with freshwater. In China, Chen et al. [117] compared a field irrigated with TWW and a non-irrigated field, finding no differences in the relative abundance of ARBs between the fields. However, the relative abundances of tet and sul genes were significantly higher in the TWW-irrigated field compared to the non-irrigated field. Mclain et al. [118] compared the soil of a basin recharged with TWW for over 20 years and the soil of a groundwater-filled basin. A higher proportion of MDR bacteria was detected in the groundwater-filled basin than in soil exposed to TWW. Troiano et al. [36] compared soils irrigated with treated GW from constructed wetlands to soils irrigated with freshwater, with no difference in the abundance of tetracycline-resistant bacteria between the two types of soils. Cerqueira et al. [120] investigated a field irrigated with river water receiving wastewater (up to 92% of the water flow was wastewater effluent). The relative abundance of mecA, tetM, and blaOXA-58 genes was higher in the field with TWW irrigation, while sul1 and blaTEM remained unaffected. Commercial agriculture fields irrigated with TWW were studied by Marano et al. [121] compared to fields irrigated with either surface water, groundwater, or desalinated water. However, even with TWW irrigation, almost all ARG concentrations were below detection limits in all the tested soils. They found no correlation between ARG abundance in irrigation water and those detected in soil. Kampouris et al. [109] compared ARGs in fields irrigated with TWW during periods of different irrigation intensity and a no-irrigation period. The relative abundance of tet, bla, sul, and intl1 genes was higher with intensive irrigation than during the period with no irrigation. The relative abundance of several ARGs increased as the intensity of irrigation increased. Freshwater irrigation and TWW irrigation were compared by Marano et al. [122] in a laboratory setting. In freshwater-irrigated soil, Gram-positive bacteria, Bacillaceae, vigorously proliferated, while TWW irrigation stimulated the growth of Enterobacteriaceae and Moraxellaceae families.

Based on currently available data, the high number of ARGs and ARBs in TWW cannot thrive or survive in the soil environment. Thus, they do not significantly contribute to antibiotic resistance of the soil microbiome [123]. There is no selective pressure significant enough from TWW effluents to induce the propagation of ARGs in soils. This suggests that the impact of ARGs on the soil microbiome is negligible and that the high levels of ARBs and ARGs in both the freshwater- and the TWW-irrigated soils indicate native antibiotic resistance in the natural soil microbiome. TWW irrigation might induce changes in microbial activity and community composition primarily because of its impacts on soil properties in different ways, such as increased salinity and dissolved organic matter levels [124].

4. Irrigation Spiked with EMPs

The presence of EMPs significantly impacts the microbial community structure in GW. Biological processes are disrupted by decreased metabolism, affecting the ability of microorganisms to treat GW [125]. EMPs are removed in the treatment system by substrate adsorption and biodegradation, the latter playing the most critical role [126,127]. However, different GW treatment strategies can either promote or remove ARGs [128]. This makes understanding the dynamics between EMPs and ARG in GW treatment systems even more complex. For example, Lai et al. [129] reported that ARGs were detected in the biofilm, having 5–35 times higher concentrations than the treated wastewater effluent. Zhou et al. [104] proposed that a change in dissolved oxygen concentrations led to significant differences in ARG levels and changed the microbial community. An increased dissolved oxygen concentration led to fewer ARG accumulation in the biofilm and a higher abundance of aerobic bacteria.

The introduction of EMPs into the soil and its effect on ARGs in the soil microbiome is yet to be studied. The accumulation of EMPs in the soil leads to an imbalance of microorganisms and a reduction of agricultural production efficiency [130]. This is because the soil microbiome can reduce stress factors, such as salinity, heavy metals, and drought, improving plant growth [131]. Some studies have indicated that some EMPs in agricultural soil can reduce the diversity of the microbial community and completely change its structure [132]. Considering the high variation and complexity of soil microbial communities from various locations [133,134], more studies are needed to cover the knowledge gap regarding the influence of treated GW irrigation on ARG prevalence in soils and soil microbial communities. The currently available knowledge is summarized in

Table 4. The effect of EMPs on the soil microbiome.

Micropollutant Exposure	Methodology	Effect Exposure	Reference
Triclosan	Soil microcosms irrigated with synthetic GW, supplemented with 2.0 μg/mL TCS	- ARB amount increased - microbial diversity decreased	[135]
Triclosan	Biolog ECO plates were used with soil from a rice paddy. The soil was spiked with six different concentrations (0, 0.1, 1, 10, 30, 50 mg/kg), sealed inside a plastic bag, incubated at 28 °C in the dark, and read every 24 h over 7 days.	- no adverse effects on the functional diversity of the soil microbial community - Shannon’s diversity index increased	[136]
A mixture of 14 pharmaceuticals	Irrigation with wastewater spiked at 10 and 100 μg/L in a controlled greenhouse experiment	- no change in the abundance of bacterial communities - no change in the abundance of ammonium-oxidizing bacteria - no impact on sulfamethoxazole-resistant and -degrading bacteria - limited effect on the bacterial evenness	[137]
Triclosan	Soil exposed to TCS at concentrations of 0, 1, 5, 10, 50, and 100 mg/kg of soil	- minimal effect on enzyme activity - no effect on general microbial activity	[138]
Triclosan	Freshwater microbes exposed to TCS, alone or together with three different micropollutants	- Planctomycetes and Actinobacteria (more sensitive bacteria) were replaced by Gammaproteobacteria and Bacteroidetes (more tolerant)	[139]
PPCPs	Soil irrigated long-term with wastewater compared to soil irrigation with groundwater	- significant reduction of bacterial abundance	[130]
Triclosan	Soil microcosms exposed to different concentrations of TCS	- significant reduction in soil microbial biomass	[140]

.

Triclosan is the most studied chemical, and several studies have shown its effects on the terrestrial microbial community. TCS will likely accumulate in the soil because of its low leaching potential [141]. In the environment, TCS can function as a toxic component and a carbon source [142]. Harrow et al. [135] documented that while there was no change in heterotrophic microorganisms in the soil, the kind of microorganisms and their antibiotic resistance were significantly influenced by triclosan. The number of bacteria resistant to antibiotics increased. At the same time, the overall diversity of the microbial community decreased, while it did not show a significant change in soil irrigated with synthetic greywater without triclosan. Research by Liu et al. [136] showed that adding TCS to soil did not negatively affect the functionality and diversity of the soil microbial community. However, it was also demonstrated by Waller and Kookana [138] that the nitrogen cycle could be disturbed by TCS at concentrations below 10 mg/kg through its impact on soil microorganisms. Izabel-Shen et al. [139] exposed soil microcosms to TCS either alone or in a mixture with other EMPs. The microcosms received a fresh medium with the EMPs every 5 days, while the change in the microbial community was monitored. They documented that more sensitive bacterial species were replaced by more resistant species, such as Gammaproteobacteria. Research by Zaayman et al. [140] showed that exposing soil microcosms to different concentrations of TCS changed several soil characteristics; some of them were more sensitive than others. Therefore, in order to measure soil health, more than one parameter should be measured. A significant reduction in soil microbial biomass has been detected for all levels of TCS exposure. TWW-irrigated soils have been found to have residual levels of EMPs, but currently available methodologies cannot link these EMPs to the soil microbial activity [137].

5. Discussion

With the rapidly increasing use of antimicrobial products and disinfectants, it is critical to evaluate the emergence of ARGs, ARBs, and EMPs in treated GW and its fate in the environment. In this review, it has been demonstrated that recent research has proven that non-antibiotic chemicals, such as disinfectants and non-antibiotic pharmaceuticals, can accelerate the dissemination of ARGs. The role of antibiotics in promoting ARBs might have been previously overestimated. Therefore, it is urgent to evaluate the role and mechanisms of non-antibiotic chemicals in spreading ARGs in the environment.

The impact of EMPs on the increase of ARGs and ARBs is challenging to estimate, especially when antibiotic resistance is not always stable and can often be reversed [143]. Another point of complexity is when the development of resistance to one micropollutant subsequently leads to the resistance to one or more micropollutants, also called cross-resistance mechanisms. Therefore, an analysis of the influence of single compounds and a mixture of these contaminants is also needed. It must be discovered when a mixture of chemicals equals the sum of particular effects, or if one or more chemicals change the interaction. So far, most experiments have focused on the impact of a single compound under controlled or laboratory conditions within a limited timeframe. Only a few report the fate and the effect of complex mixtures of contaminants in the natural environment or field conditions. Sorption to environmental particles or experimental materials can influence the impact of EMP [117]. Therefore, the effect of these preservatives needs to be analyzed in situ and in vivo, such as towards terrestrial microbial communities as a whole. In addition, more studies need to focus on identifying the mechanism/s involved in the degradation of non-antibiotic chemicals and the presence of catabolic pathways in the metagenome.

The health of humans and animals is indistinguishably connected to the environment since they can acquire AMRs from the environment through food, water, and air. Even non-pathogenic bacteria in the environment can pose a health risk, as they may function as a reservoir of ARGs and transfer it to pathogenic bacteria [144]. In addition, we tried to summarize the effects of EMPs on bacteria and their fate in the soil microbiome. EMPs contaminants have been shown to induce changes in the prevalence and diversity of soil bacteria. However, it should be considered that the fate and transport of EMPs, ARBs, and ARGs can vary with varying soil properties [145].

Some ARGs might have a different physiological role than antibiotic resistance, supported by increasing evidence [146]. Shi et al. [124] suggest that ARGs have alternative functions other than antibiotic resistance; for instance, they have biosynthetic, metabolic, or signaling purposes. For example, the MDR pumps export a wide range of different elements and have a function in bacterial chemical communication (quorum sensing) [147]. Another example is chromosomal β -lactamases in Enterobacteriaceae that contribute the resistance of β -lactam antibiotics [148]. Enterobacteriaceae share a common ancestor, and chromosomal β -lactamases are present in all species. This indicates that chromosomal β -lactamases were acquired before this genus differentiated evolutionarily into species, which happened hundreds of thousands of years ago [149]. Thus, the activity of chromosomal β -lactamases against antibiotics can be considered a side effect, and they are mainly enzymes involved in synthesizing peptidoglycans [150]. However, once antibiotic selective pressure is applied, it can reinforce their adaptive role as antibiotic resistance factors. In summary, the fact that ARGs have alternative functions supports the idea that antibiotic resistance should be examined more holistically.

Stanton et al. [151] reported that the dissemination of AMR from the environment to humans is a subject that needs more attention. Several countries have planned strategies to combat antibiotic-resistant pathogens, including reducing the unnecessary release of antibiotics into the environment [152], improving international collaboration, and supporting research for next-generation antibiotics. These policies have efficiently reduced antibiotic resistance in developed countries, but antibiotic resistance is still increasing in developing countries [144]. To significantly reduce antibiotic resistance, we should reduce the high load of antimicrobials in the wastewater. We should consider redesigning wastewater treatment in which EMPs are either degraded or removed during the process. The removal and degradation of EMPs from wastewater require advanced treatment techniques [153].

6. Conclusions

This review revealed that ARGs, ARBs, and emerging pollutants are present in GW, both raw and treated. EMPs have a non-target-specific effect; therefore, bacterial resistance can be developed by specific protective mechanisms against antibiotics. The effect of treated GW irrigation containing ARGs and ARBs showed a variable effect on the soil microbiome. In addition, limited data showed that emerging pollutants in treated GW irrigation had different impact on ARBs in the soil and microbial diversity. Exposing soil microcosms to various EMPs changed several soil characteristics, emphasizing the need to determine the effects on soil health more holistically.