key: cord-0938608-lgg8ikr5 authors: Abbott, Timothy; Kor-Bicakci, Gokce; Islam, Mohammad S.; Eskicioglu, Cigdem title: A Review on the Fate of Legacy and Alternative Antimicrobials and Their Metabolites during Wastewater and Sludge Treatment date: 2020-12-03 journal: Int J Mol Sci DOI: 10.3390/ijms21239241 sha: 3f1a005392bc91913d23632d99704cbdcca1f5ea doc_id: 938608 cord_uid: lgg8ikr5 Antimicrobial compounds are used in a broad range of personal care, consumer and healthcare products and are frequently encountered in modern life. The use of these compounds is being reexamined as their safety, effectiveness and necessity are increasingly being questioned by regulators and consumers alike. Wastewater often contains significant amounts of these chemicals, much of which ends up being released into the environment as existing wastewater and sludge treatment processes are simply not designed to treat many of these contaminants. Furthermore, many biotic and abiotic processes during wastewater treatment can generate significant quantities of potentially toxic and persistent antimicrobial metabolites and byproducts, many of which may be even more concerning than their parent antimicrobials. This review article explores the occurrence and fate of two of the most common legacy antimicrobials, triclosan and triclocarban, their metabolites/byproducts during wastewater and sludge treatment and their potential impacts on the environment. This article also explores the fate and transformation of emerging alternative antimicrobials and addresses some of the growing concerns regarding these compounds. This is becoming increasingly important as consumers and regulators alike shift away from legacy antimicrobials to alternative chemicals which may have similar environmental and human health concerns. Emerging contaminants consist of wide-ranging and expanding classes of anthropogenic and naturally occurring unregulated substances, on the occurrence, behavior, environmental fate and toxicity of which there are limited data [1] [2] [3] . They include pharmaceutical and personal care products (PPCPs), steroid hormones, brominated flame retardants, surfactants, pesticides, herbicides, perfluorinated compounds, nanoparticles and many other emerging compounds that are commonly found in municipal, agricultural and industrial wastewater [1] [2] [3] [4] [5] . These chemicals are potentially suspected to have adverse effects on aquatic environments, terrestrial organisms and humans [6] . Wastewater treatment plants (WWTPs) are the major transfer route for emerging contaminants to be released into the environment, in addition to many other minor routes [4, 5] . Municipal WWTPs are not capable of fully treating these pollutants, as current facilities are generally designed to remove readily/moderately Occurrence of different classes of pharmaceuticals and personal care products in primary, secondary and digested sludges (the number of compounds for each class is shown in brackets). Adapted from reference [23] . Increasing concerns have been raised regarding endocrine-disrupting antimicrobials because of their potential human and environmental health threats and their physicochemical characteristics [31] . Antimicrobial compounds, which are either natural or semi/synthetic in origin, kill or inhibit many harmful organisms [32] . A wide range of synthetic antimicrobial agents are used in pharmaceuticals, building materials, personal care products and household consumables [29] . Specifically, TCS and TCC are commonly used antimicrobial agents with fungicidal and bactericidal properties for killing microorganisms indiscriminately, rapidly and by non-specific action [31, 33] . Alternative compounds including quaternary ammoniums and other antimicrobials are becoming increasingly common due to shifting consumer preferences and bans on TCS and TCC in Europe, the U.S. and elsewhere. Little is known regarding the presence, concentrations, fate and metabolites of these alternative compounds, which may present similar challenges to the legacy antimicrobials TCS and TCC. Even when used as they are intended, significant quantities of antimicrobials end up in wastewater, which are conveyed to wastewater treatment facilities in sewage [31] . Antimicrobials in fabrics, plastics and other durable goods along with those in soaps, cosmetics, mouthwashes and other products intended for direct contact with skin or mucosa can enter wastewater. Ingested or absorbed antimicrobials can be eliminated in humans and animals as their untransformed parent compounds, or they can be transformed into a variety of conjugated forms and/or into secondary metabolites [34] [35] [36] [37] [38] . Figure 2 represents the potential pathways for releases of legacy and emerging alternative antimicrobials and their metabolites from products to the environment. The fate of antimicrobials during wastewater and sludge treatment can be determined by investigating their concentrations in both aqueous (i.e., wastewater) and solid (i.e., wastewater treatment sludge) matrices. Significant antimicrobial removals from the aqueous phase have been observed during wastewater treatment. For example, TCS removals of 91 to 98% have been reported during conventional activated sludge treatments [8, [39] [40] [41] [42] , resulting in low TCS concentrations (0.11 to 10.7 µg/L) in treated effluent discharged to surface waters [8, [39] [40] [41] [42] . Similar results have been reported for TCC, which has similar chemical properties and usage patterns [23] . Despite apparent antimicrobial removals in the aqueous phase during wastewater treatment, only a portion of the antimicrobial removals reported in the literature are a result of biodegradation or mineralization. Many antimicrobials present in wastewater preferentially partition into particulate organic matter/sludge during wastewater treatment [8, 31, 39, [42] [43] [44] due to their low water solubilities and high (≥4) logarithmic octanol-water partitioning coefficients (Log K ow ). This typically results in substantially higher antimicrobial concentrations in wastewater sludge versus wastewater influent. Furthermore, many sludge stabilization treatments, including anaerobic digestion, degrade organic matter at a significantly higher rate than that of antimicrobials, resulting in higher antimicrobial concentrations in digested versus undigested sludge (in terms of mass analyte per gram of dry solids). Contaminants remaining in stabilized sludge may then enter the environment through land application of biosolids through soils [41, [45] [46] [47] , uptake in plants and crops [48] [49] [50] [51] [52] [53] and movement through run-off or through groundwater [54, 55] . The fate of antimicrobials during wastewater and sludge treatment can be determined by investigating their concentrations in both aqueous (i.e., wastewater) and solid (i.e., wastewater treatment sludge) matrices. Significant antimicrobial removals from the aqueous phase have been observed during wastewater treatment. For example, TCS removals of 91 to 98% have been reported during conventional activated sludge treatments [8, [39] [40] [41] [42] , resulting in low TCS concentrations (0.11 to 10.7 µ g/L) in treated effluent discharged to surface waters [8, [39] [40] [41] [42] . Similar results have been reported for TCC, which has similar chemical properties and usage patterns [23] . Despite apparent antimicrobial removals in the aqueous phase during wastewater treatment, only a portion of the antimicrobial removals reported in the literature are a result of biodegradation or mineralization. Many antimicrobials present in wastewater preferentially partition into particulate organic matter/sludge during wastewater treatment [8, 31, 39, [42] [43] [44] due to their low water solubilities and high (≥4) logarithmic octanol-water partitioning coefficients (Log Kow). This typically results in substantially higher antimicrobial concentrations in wastewater sludge versus wastewater influent. Furthermore, many sludge stabilization treatments, including anaerobic digestion, degrade organic matter at a significantly higher rate than that of antimicrobials, resulting in higher antimicrobial concentrations in digested versus undigested sludge (in terms of mass analyte per gram of dry solids). Contaminants remaining in stabilized sludge may then enter the environment through land application of biosolids through soils [41, [45] [46] [47] , uptake in plants and crops [48] [49] [50] [51] [52] [53] and movement through run-off or through groundwater [54, 55] . A significant fraction of antimicrobials removed during wastewater treatment and/or sludge stabilization are often not completely mineralized [42, 56] . Rather, they often form various conjugates and intermediate microbial metabolites, some of which be converted back into their parent compounds while others are even more toxic and biopersistant than their parent antimicrobials A significant fraction of antimicrobials removed during wastewater treatment and/or sludge stabilization are often not completely mineralized [42, 56] . Rather, they often form various conjugates and intermediate microbial metabolites, some of which be converted back into their parent compounds while others are even more toxic and biopersistant than their parent antimicrobials [7, 57, 58] . Once released to the environment, some of these byproducts can be transformed back into their parent antimicrobials, can accumulate in various environmental compartments or they can be further degraded. Therefore, it is vital that any examination of the fate of antimicrobials in wastewater and sludge treatments also considers the fate of these compounds and their partial degradation products. Triclosan (5-chloro-2-(2,4-dichlorophenoxy)phenol), chemical abstract service (CAS) #3380-34-5, is incorporated into numerous consumer and healthcare products in many countries around the world. TCS is a synthetic, nonionic, lipid-soluble broad-spectrum antimicrobial which is effective against most Gram-negative and Gram-positive bacteria [59, 60] , against some fungal species [61, 62] and against some parasites and protozoa [63] . At low concentrations (≥0.1 mg/L), it shows bacteriostatic activity, while at higher concentrations (≥2.0 mg/L), it is bactericidal against many species, including some staphylococci, streptococci, mycobacteria, Escherichia coli (E. coli) and methicillin-resistant Staphylococcus aureus (MRSA) [63] [64] [65] . TCS rapidly permeates cell walls and disrupts multiple cellular processes, Figure 3 . Number of scientific documents on the topic of "triclosan" published for the time period from 1975 to 2019 (data generated from reference [74] ). Table 1 , TCS is a halogenated biphenyl ether which has a similar structure and properties to several potent toxins, including dioxins, polychlorinated biphenyls and other persistent environmental contaminants [31] . TCS is highly lipophilic as it has a very high (≥4) Log Kow value, is highly insoluble in water and is relatively non-volatile [75, 76] . Due to its incomplete removal during wastewater treatment and sludge stabilization processes, TCS is consistently found throughout the environment in waters, soils, sediments and biota which receive treated wastewater effluent and/or stabilized wastewater residuals [5, 31, 39, 42, 43, [77] [78] [79] . As shown in Table 1 , TCS is a halogenated biphenyl ether which has a similar structure and properties to several potent toxins, including dioxins, polychlorinated biphenyls and other persistent environmental contaminants [31] . TCS is highly lipophilic as it has a very high (≥4) Log K ow value, is highly insoluble in water and is relatively non-volatile [75, 76] . Due to its incomplete removal during wastewater treatment and sludge stabilization processes, TCS is consistently found throughout the environment in waters, soils, sediments and biota which receive treated wastewater effluent and/or stabilized wastewater residuals [5, 31, 39, 42, 43, [77] [78] [79] . Triclosan can be transformed into various other compounds by biotic and abiotic processes and mechanisms within organisms, WWTPs and/or the environment. These can include various conjugated forms, photolysis products, chlorinated derivatives and partial mineralization metabolites, as shown in Figure 4 . Some commonly detected TCS-related compounds are the TCS conjugates methyl triclosan (MeTCS), glucuronidated triclosan (TCS-gluc) and triclosan sulfate (TCS-SO 4 ), several of which can have been shown to revert back into TCS [46, 80, 81] . About five percent of TCS is transformed into MeTCS through biomethylation of the TCS hydroxyl group during aerobic and anoxic processes in WWTPs [40, 70] . The conjugate TCS-gluc ( Figure 4 ) has been shown to be produced within organisms and has been found in animal or human blood, liver and intestines [46, 82, 83] . Chen et al. [57] observed that TCS was transformed into TCS-SO 4 through sulfation during a batch simulated aerobic activated sludge process. This agrees with a later batch study by Trenholm et al. [84] , who observed the rapid formation of TCS-SO 4 under aerobic conditions and slower formation of TCS-SO 4 under anoxic conditions using biological nutrient removal (BNR) sludge. Although Chen et al. [57] did not observe any decrease in TCS-SO 4 once it was formed over 10 days, Trenholm et al. [84] observed that TCS-SO 4 levels began to decrease after 5 days under aerobic conditions and observed the formation of a secondary TCS-SO 4 metabolite, whose mass matched 3,5-dichloro-2-hydroxy-benzene sulfonate (or a closely related isomer thereof). TCS conjugates can also have significantly different physicochemical properties that those of its parent antimicrobial. For example, MeTCS is more persistent and bioaccumulative due to its even higher Log K ow walue (5.0-5.1) and lower water solubility (0.4 mg/L) than those of TCS [46, 85] . However, both TCS-SO 4 and TCS-gluc are less lipophilic (estimated Log K ow values of 2.18 and 2.82, respectively), are far more polar and are more soluble in water (65.17 and 11.58 mg/L at 25 • C, respectively) [82, 86] . Triclosan can also form various intermediate metabolites, as shown in Figure 4 , as many biotic processes have been shown to degrade it. Microalgae and many isolated pure-culture bacterial strains have been shown to partially or completely mineralize TCS. Chlorella pyrenoidosa, Desmodesmus sp. and Scenedesmus sp. have been shown to remove between 77.2 to 99.7% of TCS through reductive dechlorination, ether bond cleavage, hydroxylation or the production of TCS conjugates through methylation and/or glucosylation [87, 88] . Hay et al. [89] observed partial TCS mineralization by an auxotrophic Sphingomonas-like strain isolated from an activated sludge basin. They observed the partial mineralization of the dichloro ring, with approximately 35% of 14 C-TCS-labelled carbon being transformed to 14 CO 2 in a complex medium. The Sphingomonas sp. strain YL-JM2C, isolated from activated sludge, was shown to completely mineralize up to 99% of TCS within 72 h at an optimum pH of 7.0 and temperature of 30 • C and formed 2,4-dichlorophenol (2,4-DCP), 2-chlorohydroquinone and hydroquinone as intermediate decomposition products [90] . Mulla et al. [91] proposed that YL-JM2C first oxidatively transformed TCS into 2,4-DCP by breaking the ether bond between the two aromatic rings, likely facilitated by the enzymes monooxygenase and/or dioxygenase. Then, 2,4-DCP was reductively dechlorinated into 2-chlorohydroquinone and hydroquinone, before being fully mineralized into carbon dioxide. Sphingomonas sp. PH-07, originally isolated from activated sludge by Kim et al. [92] , was shown to catabolize TCS into various hydroxylated metabolites, including monohydroxy-TCS and dihydroxy-TCS, and also formed 4-chlorophenol and 2,4-DCP, suggesting both aromatic rings underwent dihydroxylation [93] . Similar pathways were observed in an isolated Sphingopyxis strain KCY1, which was shown to rapidly degrade up to 90% of TCS within 24 h and showed 100% removal within 48 h [94] . Two oxygenase-expressing bacteria Mycobacterium vaccae JOB5 and Rhodococcus jostii RHA1 were able to degrade TCS through a meta-cleavage pathway when grown on propane and biphenyl, respectively [95] . Biphenyl-grown RHA1 produced 2,4-DCP, mono and dihydroxy-TCS and 2-chlorohydroquinone (ibid). . Common pathways for the transformation and degradation of triclosan. The primary reactions above are identified as follows: [1] glucuronidation; [2] methylation; [3] sulfation; [4] hydroxylation; [5] meta-cleavage of C-O bond; [6] chlorine substitution; [7] photolysis; [8] reductive dechlorination. Common pathways for the transformation and degradation of triclosan. The primary reactions above are identified as follows: [1] glucuronidation; [2] methylation; [3] sulfation; [4] hydroxylation; [5] meta-cleavage of C-O bond; [6] chlorine substitution; [7] photolysis; [8] reductive dechlorination. Mixed microbial cultures under methanogenic, denitrifying (anoxic) and aerobic conditions have also been shown to be effective in degrading TCS. For example, Gangadharan et al. [96] observed TCS removals of 87% and 80% during a 10-day solids retention time (SRT) when acetate was used as a co-substrate under methanogenic conditions and under denitrifying conditions, respectively, forming 2,4,DCP, phenol, catechol and other unidentified products under both conditions. Chen et al. [57] observed the formation of 2,4-DCP and 4-chlorocatechol (4-CC) in a 10-day batch simulated activated sludge reactor spiked with high levels of TCS which was formed through the meta-cleavage of the C-O bond between the two phenol rings. They also observed the production of monohydroxy-triclosan and dihydroxy-triclosan through the partial hydroxylation of the OH-containing TCS ring. Other TCS transformation routes include abiotic processes, including chemical and photochemical reactions. However, many of these degradation products are far more toxic and biopersistant than the parent TCS. For example, Noutsopoulos et al. [97] found 92% TCS removal at a free chlorine dose of 40 mg Cl 2 /L and at a contact time of less than 5 min, but observed a 50-to 100-fold increase in Vibrio fischeri toxicity compared to TCS prior to chlorination. Multiple TCS-containing personal care products have also been shown to react with free chlorine to form higher-chlorinated TCS intermediates, including dichloro-2-(2,3-dichlorohenoxy)phenol and trichloro-2-(2,3-dichlorohenoxy)phenol. This was through Cl − substitutions, various chlorophenols (2,4-DCP and 2,4,6-trichlorophenol) through the meta-cleavage of the TCS C-O bond or its higher-chlorinated intermediates and chloroform formation from ring cleavage of TCS, TCS-intermediates, chlorophenols and Cl − substitutions [98] [99] [100] . The formation of even higher-chlorinated tetrachloro-2-(2,3-dichlorohenoxy)phenol and tetrachloro-and pentachloro-hydroxyl-TCS derivates was observed by Kanetoshi et al. [101] when TCS-impregnated fabric was exposed to a 0.2% sodium hypochlorite solution in warm water (45 • C), similar to what would be used for household laundry. TCS is particularly photodegradable in its phenolate form, and photochemical processes can transform a significant amount of TCS in aquatic systems, accounting for up to 80% of TCS removal in lakes [102, 103] . Other studies have also shown that exposure of treated WWTP effluent or TCS-contaminated water can transform TCS into various higher-chlorinated TCS derivatives, dioxins and benzofurans, including dichlorodibenzo-p-dioxins (DCDDs) and 2,4-DCP, when aqueous TCS is irradiated with sunlight/UV light, even at relatively low irradiation intensities [103] [104] [105] [106] [107] . Depending on the conditions, 3 to 12% of TCS was converted to DCDDs and 2,4-DCP, respectively, with some TCS being coupled to humic matter [108] . Furthermore, the production of chloroform and other disinfection byproducts can be significantly enhanced at higher pH values or when UV is combined with chlorine treatments and/or seawater [100, 103, 109, 110 ]. Wastewater undergoes a variety of different treatments and conditions (i.e., aerobic, anaerobic, anoxic, high/low pH and/or reduction/oxidation potentials, etc.) as it is processed in a WWTP. For example, wastewater undergoes a variety of oxidation-reduction potentials in BNR processes, during which different microorganisms are active and a wide range of metabolic processes are occurring. A portion of antimicrobials can be degraded and/or transformed into other compounds during treatment. Various studies have reported TCS concentrations of 0.3 to 12.5 µg/L in domestic wastewater [41, 42, 70, [111] [112] [113] [114] . Trace amounts of MeTCS in untreated wastewater have also been reported, with values between 0.015 to 1 µg/L [41, 42, 70, 112, 113] . TCS is largely removed from the aqueous phase of wastewater during treatment, as studies that analyzed WWTP effluents have demonstrated near-complete TCS removals ranging between 85 and 98% [39] [40] [41] [42] 56, 70, [115] [116] [117] [118] . Despite high levels of removal from the aqueous phase of wastewater, only a fraction of this is due to TCS being fully or partially mineralized [31] . Numerous studies have shown that 66 to 80% of TCS is associated with particulate or solid matter in wastewater [8, 39, 42, 116] . Soluble TCS tends to partition (or sorb) to particulate matter/sludge and can accumulate to concentrations several magnitudes higher during wastewater treatment [42, 119] . This is likely due to the very high surface areas of wastewater sludge available for sorption (ranging from 0.8 to 1.7 m 2 per gram of material) and the highly hydrophobic and lipophilic nature of TCS [31, 75] . Significant TCS removals have been observed during primary settling through sorption and settling of solids, with up to 75% TCS removals being observed from the aqueous phase, although little to no biotransformation during primary treatment was observed [42] . Triclosan has been shown to degrade under aerobic conditions, and a wide range of biotransformation values during secondary (biological) treatment have been observed. Guerra et al. [111] found that TCS removal in WWTPs was strongly correlated with total Kjeldahl nitrogen removals. Some laboratory studies have shown high (75 to 99%) TCS degradation during simulated activated sludge processes [85, 120] , although full-scale studies typically find significantly lower removals. For example, Lozano et al. [42] observed 10% and 23% TCS removal during conventional activated sludge and subsequent nitrification/denitrification processes, respectively, while Heidler and Halden [39] observed just under 50% removal. This is likely due to a significant portion of TCS not being exposed to aerobic conditions due to removal during primary treatment as well as incomplete mixing or shortcutting during secondary treatment. Chen et al. [85] observed that approximately 1% of TCS in a simulated activated sludge process was transformed into MeTCS during aerobic conditions while no TCS removal or MeTCS generation occurred under simulated anoxic or anaerobic conditions. A simulated batch activated sludge process produced various monoaromatic compounds, including 4-CC and 2,4-DCP and several isomers of mono-and di-hydoxy-TCS, and approximately 7% of the initial TCS was transformed into TCS-SO 4 , which accumulated and persisted until the end of the experiment [57] . TCS removals in secondary clarifiers are primarily due to the settling of TCS-bearing biomass, and little change was reported in aqueous-phase TCS or MeTCS concentrations [39, 42] . Although several studies have found that TCS can be transformed during chlorination or through UV disinfection processes, Lozano et al. [42] did not find any significant changes in dissolved TCS, and MeTCS remained constant during filtration and disinfection processes. Typical literature values in treated effluent range from 0.07 to 6.20 µg/L for TCS and from 0.015 to 0.02 µg/L for MeTCS [41, 42, 70, 112] . Relatively high concentrations of TCS and MeTCS have been reported in primary and secondary sludge as they are the primary sinks for TCS and TCS-related compounds. A survey of six Canadian WWTPs published in 2015 found primary sludge TCS concentrations of 6 to 14 µg/g dw, while a 2009 study of a large U.S. facility reported average primary sludge TCS concentrations of 20.3 ± 0.9 µg/g dw and 0.09 ± 0.002 µg/g dw of MeTCS [42, 121] . These studies reported somewhat lower concentrations in thickened waste activated sludge, with TCS values of 0.43 to 14.14 µg/g dw and 0.25 ± 0.04 µg/g dw MeTCS in the U.S. study (ibid). Mixed primary and secondary sludge TCS concentrations reported in the literature vary widely, with average values between 4.0 and 12.9 µg/g dw in several Canadian WWTPs [43, 80, [122] [123] [124] , which were comparable to values found in the U.S. and elsewhere [125] . Other than MeTCS, few studies have found significant levels of other TCS metabolites in undigested sludge from full-scale facilities, with the exception of Abbott and Eskicioglu [80] who reported 0.1 to 0.2 µg/g of TCS-SO 4 in undigested mixed sludge from a small-to medium-scale BNR facility. Table 2 presents the occurrences of TCS and its transformation products/metabolites in sludge and in various stabilization processes. Many emerging contaminants are somewhat resistant to degradation and typically degrade at much lower rates than that of organic matter during sludge stabilization processes. Due to this, effluent from many stabilization processes, such as anaerobic digestion, often contain far higher antimicrobial concentrations (in terms of µg/g dw) and appear to be concentrated as sludge volumes that are typically reduced by 50% or more. Depending on the method of calculation, many studies have found negative TCS removals, and some refer to this as "accumulation" during digestion. For example, McAvoy et al. [41] found that anaerobic digestion was ineffective in degrading TCS in primary sludge with significant (108%) TCS accumulations in anaerobic digester digestate (15.6 µg/g dw). This study also observed the generation of small amounts of MeTCS (0.13 µg/g dw) and some (0.09 to 0.11 µg/g dw) higher-chlorinated TCS derivatives, including 2,3-dichloro-6-(2,4-dichlorophenoxy) phenol and 2,3,4-trichloro-6-(2,4-dichlorophenoxy) phenol. Heat-treated biosolid pellets from a full-scale WWTP in the United Kingdom. n/a n/a 11,220 to 28,220 ng/g dw Biosolid pellets had 35 to 69 ng/g dw of MeTCS [128] Five American WWTPs, including two activated sludge (treating wastwater from 400,000 and 27,000 people) and three tricking filter (all treating wastewater from <3100 people). Full-scale aerobic digestion and anaerobic digestion of mixed sludge. Process specifics not provided. • Activated sludge facilities: -8750 to 14,700 ng/g dw TCS in primary; -900 to 4,200 ng/g dw TCS in secondary sludges. • Tricking filter facilities: -7500 to 12,200 ng/g dw TCS in primary; -7300 ng/g dw TCS in secondary sludges. • Digested sludge values not reported in activated sludge facilities • 530 to 15,600 ng/g dw in tricking filter facilities Study also examined MeTCS and higher-chlorinated TCS compounds: • MeTCS in activated sludge facilities: -50 to 260 ng/g dw in primary -310 to 1030 ng/g dw in secondary • MeTCS in tricking filter facilities: -140 to 200 ng/g dw in primary sludge -450 ng/g dw in secondary sludge • Low levels of higher-chlorinated TCS in digested tricking filter sludge: -< LOD to 110 ng/g dw of 4,5-dichloro-2-(2,4-dichloro-phenoxy)-phenol -< LOD to 100 ng/g dw of 5,6-dichloro-2-(2,4-dichloro-phenoxy)-phenol -79%) was transferred to the biosolids. Among the 48 target PPCPs studied by Sun et al. [199] , TCC was detected in all of the influent, effluent and sludge samples (n = 12) collected from three WWTPs located in different areas in China over one year. Average TCC concentrations in the influent and effluent ranged from 4.7 to 109 ng/L, whereas TCC levels in sludge were in the range of 1130-2180 ng/g dw with an average of 1700 ng/g dw. The occurrence of TCC in aqueous and solid waste streams of seven Chinese WWTPs was investigated in another study by Wang et al. [9] . Although dissolved TCC had relatively low concentrations (between 4.78 and 500 ng/L, with detection frequencies of >95.9%) in WWTP influent and effluent samples, adsorbed TCC was detected in all suspended solids and sludge samples (n = 49), with median concentrations of 968 and 1090 ng/g dw, respectively. On the other hand, among the 19 biocides monitored in eight municipal WWTPs in Thailand, TCC was determined as the predominant compound in the influent, effluent and excess sludge, with average concentrations of 838 ng/L, 95.6 ng/L and 4010 ng/g dw, respectively [29] . These studies clearly indicate that the sorption of TCC to sludge was found to be the main mechanism in its removal during wastewater treatment. TCC has been identified as one of the top trace contaminants among wide-ranging PPCPs in sludge and biosolids assayed in several countries, such as the U.S. [18, 201, 204] , Canada [205] , China [32, 206] , Japan [117] , Korea [207] and Thailand [29] . TCC was found to be one of the most abundant analytes (48% of total PPCP mass found) among the 72 PPCPs detected in biosolids samples obtained from 94 U.S. WWTPs in 32 states and the District of Columbia, and its mean concentration was 36,060 ± 8049 ng/g dw [18] . The occurrence of TCC within 72 PPCPs assayed in sludge samples taken from 74 facilities was also investigated by the U.S. EPA in its 2009 "Targeted National Sewage Sludge Survey". TCC was detected in all of the sludge samples at concentrations ranging from 187 to 441,000 ng/g dw [125] . Moreover, the levels of TCC in lime-stabilized biosolids samples, which were taken from a WWTP in the Mid-Atlantic region of the U.S. at around every 2 months over a 7-year period, were between 8850 and 22,900 ng/g dw, with an average concentration of 14,300 ± 3710 ng/g dw (n = 31) [33] . TCC was also identified as a priority compound among the top 10 contaminants from the 231 emerging chemicals analyzed in the U.S. biosolids regarding its high occurrence and high bioaccumulation potential [5] . A study conducted by Guerra et al. [205] reported that TCC was present in every biosolid sample (n = 24) collected from six WWTPs in Canada and was one of the most abundant compounds, at concentrations from 1200 to 8900 ng/g dw, among the 62 PPCPs. Another study by Chu and Metcalfe [208] reported that TCC concentrations in samples collected from three Canadian WWTPs were between 2170 and 4820 ng/g dw in activated sludge and between 3050 and 5970 ng/g dw in biosolids. In the study by Liu et al. [32] , TCC was also found to be the predominant compound within 18 biocides analyzed in 10 WWTPs in South China and its concentration levels in excess sludge were between 1140 and 2950 ng/g dw. According to the first nationwide study in Korea performed by Subedi et al. [207] , TCC among 10 PPCPs was found in 100% of digested sludge samples taken from 40 representative WWTPs that receive domestic, industrial or mixed (both domestic and industrial) wastewaters. Their results showed that the median concentrations of TCC in digested sludge collected from domestic (2860 ng/g dw) and mixed (2380 ng/g dw) WWTPs were 160-200-times greater than that in digested sludge collected from industrial WWTPs (14.2 ng/g dw). Lehutso et al. [202] also reported TCC concentrations in raw sludge, treated sludge and biosolids ranging between 3.58 and 11.8, 1.95 and 8.31 and 2.59 and 11.9 ng/g dw, respectively, in six WWTPs of South Africa over a period of two years. As the anaerobic digestion process is a commonly used sludge stabilization option in municipal WWTPs because of its capability by generating bioenergy and biosolids [209] , several studies have focused on the occurrence and fate of persistent TCC during conventional anaerobic sludge digestion. In the study by Heidler et al. [119] , a mesophilic (35-37 • C) digester operated at an SRT of 19 days did not improve TCC transformation and resulted in an accumulation of TCC in dewatered, digested sludge up to concentrations of 51,000 ± 15,000 ng/g dw. Another study indicated the accumulation and persistence of TCC in anaerobically digested biosolids [201] . Furthermore, Yang et al. [210] observed no TCC degradation during anaerobic digestion operated under mesophilic (35.0 ± 0.5 • C) conditions at SRTs of 15, 20 and 30 days and TCC remained predominantly in the solid phase of digested sludge in their study. Kor-Bicakci et al. [44] also reported that anaerobic digesters fed with mixed sludge at different SRTs of 20, 12 and 6 days, showed limited TCC removals (from 11 to 32%) under mesophilic (35 ± 1 • C) and thermophilic (55 ± 1 • C) conditions. Narumiya et al. [117] showed that the level of TCC was reduced moderately during anaerobic digestion in a Japanese WWTP and was found at one of the highest concentrations (from 1540 to 3140 ng/g dw) in digested sludge among the 47 PPCPs examined. Consequently, it can be concluded that the transformation/degradation of TCC is very limited during the conventional anaerobic digestion. For this reason, recently, a few studies have focused on solutions (such as using sludge pretreatment technologies prior to digestion) that can boost the removal of TCC and its potential degradation products from digested biosolids. Yang et al. [211] reported that anaerobic sludge digestion using thermal pretreatment at a temperature of 150 • C for 30 min (with a pressure of 500 kPa) enhanced the biodegradation of TCC by around 10%. The impact of the Cambi Thermal Hydrolysis Process™ in conjunction with mesophilic anaerobic digestion (at an SRT of 22 days) on the levels of TCC and its transformation products in biosolids was investigated by Armstrong et al. [127] . The implementation of thermal hydrolysis, applied at temperatures of 150-180 • C for 30 min, to the full-scale WWTP allowed for a significant decrease in the levels of TCC (from 5734 ± 2182 to 630 ± 974 ng/g dw) from biosolids. In another study by Kor-Bicakci et al. [44] , advanced anaerobic digestion coupled with microwave pretreatment was found to be highly efficient in increasing TCC elimination from digested biosolids. TCC was eliminated significantly (from 43 to 78%) in the advanced digesters fed with microwave-pretreated sludge at 160 • C, while advanced digesters fed with microwave-pretreated sludge at 80 • C displayed moderately efficient TCC removal (from 17 to 47%) from digested biosolids. According to their results, within the OLR range of 1.46 to 5.25 g VS/L/day, thermophilic (55 ± 1 • C) digesters had better performances in the reduction in TCC in biosolids compared to mesophilic (35 ± 1 • C) digesters. Furthermore, Yu et al. [186] indicated that the biodegradation rate of TCC was 83.1% after composting of sludge for 16 days with forced ventilation, whereas this rate was only 18.2% after sludge stacking for 16 days. Although most studies have investigated the occurrence and fate of TCC in sludge and biosolids, only a few have reported its behavior coupled with its potential transformation products and metabolites [44, 56, 127, 201, 206] . This can be explained by the lack of accurate and cost-effective analytical methods for the quantitative analysis of complex sludge matrix and the challenges encountered during their analysis [164, 212] . Table 4 shows studies that reported the occurrences of both TCC and its transformation products/metabolites in sludge and biosolids. Sapkota et al. [213] revealed that both DCC and 4-Cl-TCC, which were previously unrecognized pollutants, were detected consistently in primary sludge samples from various WWTPs located in five different U.S. states beside TCC. In their study, high Log K ow values (4.3 and 5.6, calculated using the K ow WIN TM model) and low water solubilities (3.7 and 0.11 mg/L, calculated using the Solaris V4.67 model) were reported for DCC and 4-Cl-TCC, respectively. Based on the quantitative structure-activity relationship analyses, DCC was found to be less particle-active and 34-times more water-soluble compared to 4-Cl-TCC [213] . In another study conducted by Heidler and Halden [201] , 4-Cl-TCC and DCC were detected for the first time in anaerobically and aerobically digested sludge collected from 25 full-scale WWTPs in 18 U.S. states. In their study, TCC was found at comparatively higher levels compared to its transformation products, detected in digested sludge. Moreover, the nationwide survey by Pycke et al. [56] was the first to report TCC transformation through dechlorination occurring in the sewage collection system and within a WWTP. DCC and MCC were detected in raw and digested sludge samples obtained from 14 different WWTPs from across the U.S. The effects of varying sludge treatment processes on the levels of TCC and its dechlorination products in biosolids were also investigated in the same study [60] . Anaerobic digestion processes caused significant accumulation of these compounds in digested sludge, whereas digested sludge heat drying processes enhanced the removal of TCC, DCC and MCC. Conversely, very limited removal efficiencies for TCC and 4-Cl-TCC were obtained during the digested sludge dewatering process, while no significant changes were observed for DCC and MCC in the same process. In the study by Chen et al. [206] , TCC dechlorination during wastewater treatment was observed for the first time in China. Two human metabolites of TCC (2 -hydroxy-TCC and 3 -hydroxy-TCC) were also detected in over 90% of the 100 sewage sludge samples collected across 21 Chinese provinces/districts. In another study conducted by Kor-Bicakci et al. [44] , TCC was able to be converted into its non-chlorinated congener (NCC) during conventional and advanced anaerobic digestion. Their research was the first to report the efficient and complete TCC dechlorination occurred in advanced thermophilic and mesophilic digesters fed with microwave-pretreated mixed sludge. There are significant concerns that TCC released to the environment through wastewater sludge may have potential adverse effects on human health and on ecosystems. Due to its high chlorine content and aromatic nature, TCC exhibits a tendency for environmental persistence [45, 143, 144, 176, 214, 215] . The environmental half-life of TCC has been estimated at 60, 120 and 540 days in water, soil and sediment, respectively [176] . Laboratory experiments conducted by Ying et al. [143] showed that TCC was degraded in a loam soil with a half-life of 108 days under aerobic conditions, whereas it was highly resistant to biodegradation within 70 days of the experimental period under anaerobic conditions. The estimated half-life of TCC ranged from 87 to 231 days in soils with and without the addition of biosolids under aerobic conditions [144] . Another study also reported a soil TCC half-life of 287.5 ± 45.5 days in commercial farms that received a single biosolids application [215] . Several studies have explored potential endocrine-disrupting effects of TCC on animals (e.g., rats, frogs, zebrafish and sea urchins) [171, [216] [217] [218] [219] and humans [175, 220, 221] . TCC has also been shown to have detrimental impacts on aquatic organisms (e.g., algae, crustaceans and fish) [3, 140, 222] and on microorganisms [223, 224] . In a study by Chalew and Halden [140] , crustacea were identified as one of the most sensitive species to TCC, with acute and chronic toxicity threshold values between 0.0019 and 0.04 and 0.00006 and 0.0047 mg/L, respectively. In the same study, acute and chronic threshold levels in fish were observed in the range of 0.049-0.18 and 0.005 mg/L, respectively. According to the results of acute toxicity tests of TCC on eight Chinese aquatic species, Daphnia magna (planktonic crustacean) with a 48-h LC50 of 0.006896 mg/L was the most sensitive species to TCC, followed by Rana chensinensis (0.02384 mg/L, amphibian), Gobiocypris rarus (0,1103 mg/L, fish) and Macrobrachium nipponense (0.2616 mg/L, abenthic crustacean [222] ). TCC also has acute toxic effects with a 48-h LC50 of 0.01, 0.0031 and 0.015 mg/L for Daphnia magna, Ceriodaphnia dubia and Mysidopsis bahia (invertebrates), respectively [174] . Moreover, the 96-h LC50 values of TCC to Salmo gairdneri and Lepomis macrochirus (fish) were 0.12 and 0.097 mg/L, respectively [174] . TCC has also been shown to bioaccumulate in terrestrial organisms (e.g., earthworms) inhabiting biosolid-amended soil ecosystems [14, [225] [226] [227] [228] . On the other hand, it has been found to accumulate in organisms (e.g., snails, freshwater worms, fish and algae) living in aquatic environments exposed to effluent from WWTPs [229] [230] [231] [232] [233] . Transfer of TCC from biosolid-amended soils and treated wastewater into aquatic and terrestrial ecosystems opens a pathway for potential bioaccumulation, bioconcentration and biomagnification of TCC throughout the food chain [30] . The study carried out by Sherburne et al. [226] demonstrated that biosolids-derived TCC had been found at different trophic levels within a terrestrial food web, including earthworms (primary consumer), deer mice (secondary consumer), European starlings (a secondary consumer of invertebrates) and American kestrels (a tertiary consumer of rodents, small birds and invertebrates). Furthermore, TCC has been found to bioaccumulate in plants (e.g., carrot, green pepper, tomato, cucumber, pumpkin, zucchini and switch grass) grown in soils amended with biosolids [28, 48, 52, 53] . TCC was also found to be concentrated in root tissues and translocated into above-ground parts, such as stems, leaves and fruits [51, 53, 234] . Upon human exposure, TCC can be metabolized and excreted by the human body during its utilization in consumer and personal care products [34, 35, 235, 236] . Schebb et al. [35] reported the detection of TCC in whole blood after exposing human subjects by a single shower with antibacterial soap. TCC has also been detected in many human matrices, including urine [34, 56, 172, 237, 238] , serum [172] , breast milk [33] and maternal urine and cord blood plasma [155, 236, 239] , and in finger and toenails [237] . Ye et al. [172] measured the total and free concentrations of TCC and its phase-I metabolites (2'-hydroxy-TCC and 3 -hydroxy-TCC) in 50 urine and 16 serum samples collected anonymously from adult volunteers in the U.S. Total (free plus conjugated) TCC was detected in 5.4-28% of the urine samples together with its phase-I metabolites, whereas total TCC was detected in about 50% of the serum samples. Pycke et al. [155] also investigated the total concentrations of TCC and its human metabolites in maternal urine from 181 pregnant women in an urban population from Brooklyn, New York. Excreted TCC was detected in the urine of 86.7% of the study participants, while total 2 -OH-TC and 3 -OH-TCC were detected after conjugate hydrolysis in 27.1% and 16.6% of the urine samples, respectively. Regarding transformation/degradation products of TCC, their potential ecological and human health risks are still rarely known [162, 163, 187] . TCC is known to be precursor of toxic and carcinogenic chloroanilines in the form of mono-and di-chloroaniline [31, 192, 201] . Specifically, 3,4-DCA has been recognized as an endocrine-disrupting chemical [240] . Furthermore, 4-CA has also been known to cause cancer [241] . Because of their recalcitrant properties, 4-CA and 3,4-DCA have been ubiquitously present in the environment and have been shown to accumulate in sludge, soil and sediments [179, [242] [243] [244] . In a study by Sipahutar et al. [183] , the toxicity of TCC and its degradative intermediates (3,4-DCA and aniline) were investigated using cytogenotoxicity assessments. According to an Allium cepa root-tip chromosomal aberration assay, root-tip cells exposed to TCC at various concentrations had more chromosomal abnormalities compared to root-tip cells exposed to TCC's degradative intermediates. In addition to 4-chloroisocyanatobenzene and 4-chloronitrobenzene, 3,4-DCA and 4-CA were also identified as the main degradation products of TCC during photodegradation in aqueous systems under ultraviolet radiation [245] and simulated sunlight irradiation [246] . Concerning dechlorination products of TCC, DCC and 4-Cl-TCC are known as manufacturing byproducts in technical-grade TCC (at an estimated level of 0.2% by weight) [213] ; thus, TCC-containing personal care products can be considered as another source of the release of these compounds into the environment [177] . The higher-and lower-chlorinated congeners of TCC have been reported to be present in sediments and surface waters impacted by WWTP effluent as well as in sediment-dwelling organisms, such as the freshwater oligochaete Lumbriculus variegatus (worm) [162, 192, 193, 213, 230, 247] . A previous study conducted by Miller et al. [192] reported that the TCC dechlorination process was observed exclusively in strictly anaerobic environments (such as dated estuarine sediments) by strict anaerobic dechlorinating microorganisms. However, an alternate TCC dechlorination pathway was proposed in recent studies that took place in potentially partially oxygenated environments, such as freshwater sediments [162, 195] . Furthermore, its carbanilide congeners have also been shown to act as endocrine disruptors [171] . Based on the few studies available in the literature, further research is required to better understand the potential impacts of these transformation/degradation products. Products containing emerging alternative antimicrobials are becoming increasingly common in the market place [129, [248] [249] [250] . Most hand sanitizers in the present U.S. market contain emerging antimicrobials [251] , while the recent novel coronavirus pandemic has also created an increasing demand for hand sanitizers and other disinfection products. This has created a huge market opportunity for manufacturers to continue to produce and market large amounts of antimicrobial-containing products [252] . Therefore, it is vitally important to investigate any potential environmental impacts and human health effects of these alternative antimicrobials, as the excessive use of them may lead to similar environmental and human health concerns as legacy antimicrobials. More than 4500 scientific papers have been published on the topic of "benzalkonium chloride" between 1975 and 2019 (Figure 7) . Over the last decade, benzyl-substituted quaternary ammonium chlorides (QACs) have become common in the marketplace as emerging alternative antimicrobials. The history of the use of QACs as a cleaning agent dates back to 1935, when Domagk reported the enhanced efficacy of the substituted long-chain alkyl groups [253] . In the beginning, quaternary ammonium salts were thought to be innocuous and green chemicals which possessed powerful germicidal and antimicrobial activities. Consequently, the extensive application of BACs in product formulations has resulted in the wide distribution of them in the environment, particularly in wastewater and sludge [254] [255] [256] [257] . This section will focus on the environmental fate and distribution of BACs since they are one of the few classes of compounds which are currently not restricted and are being used Over the last decade, benzyl-substituted quaternary ammonium chlorides (QACs) have become common in the marketplace as emerging alternative antimicrobials. The history of the use of QACs as a cleaning agent dates back to 1935, when Domagk reported the enhanced efficacy of the substituted long-chain alkyl groups [253] . In the beginning, quaternary ammonium salts were thought to be innocuous and green chemicals which possessed powerful germicidal and antimicrobial activities. Consequently, the extensive application of BACs in product formulations has resulted in the wide distribution of them in the environment, particularly in wastewater and sludge [254] [255] [256] [257] . This section will focus on the environmental fate and distribution of BACs since they are one of the few classes of compounds which are currently not restricted and are being used extensively in the market in a wide range of products [248] . Benzalkonium chlorides are the most widely used compounds currently deferred to as an eligible active ingredient for consumer antiseptic products in the final rule of the 2019 U.S. FDA Federal Register [248] . Due to the extensive use of BACs as cationic surfactants, phase transfer agents, fabric softeners, disinfectants, detergents, biocides and as ingredients in various personal care and pharmaceutical preparations, BACs have been detected in environmental samples of surface water, runoff water, sediment, wastewater and sewage sludge [254] [255] [256] [258] [259] [260] [261] [262] [263] [264] [265] . The most commonly used BACs are alkyldimethylbenzylammonium chlorides, in which the quaternary ammonium ion contains two methyl groups and one benzyl group on one side and a long-chain alkyl group (with typical even-numbered carbon atoms C 8 -C 18 ) on the other side. BACs are a broad class of homologous n-alkylbenzyldimethylammonium chlorides with varying numbers of carbon atoms in the alkyl chain. Commercial formulations use either one or a mixture of homologues in order to meet the selective strength and activity of applications. The strength is influenced by the length of the alkyl chain [266] . The most commonly used commercial products contain BAC-12, BAC-14 and BAC-16, referring to 12, 14 and 16 carbon atoms in the alkyl chain, respectively [267, 268] . The hydrophobic alkyl end plays a contributory role to the adsorption characteristics of these compounds and, hence, their fate, partitioning and distribution in the environment [254, 269, 270] . The physicochemical properties of commonly used BACs are shown in Table 5 . The biodegradation pathways for BACs in wastewater and sludge have not yet been well established. Some research papers have shown degradation mechanisms along with their byproducts and metabolites in the wastewater treatment systems during chemical oxidation processes [273] [274] [275] [276] . In water-treatment systems, the formation of disinfection byproducts (DBPs), particularly N-nitrosodimethylamine (NDMA), has been well studied due to the carcinogenic characteristics of NDMA [276] [277] [278] [279] [280] [281] . BACs contain a quaternary nitrogen atom which can serve as a possible precursor for the formation of DBPs [282] [283] [284] . BACs are used as algaecides in swimming The biodegradation pathways for BACs in wastewater and sludge have not yet been well established. Some research papers have shown degradation mechanisms along with their byproducts and metabolites in the wastewater treatment systems during chemical oxidation processes [273] [274] [275] [276] . In water-treatment systems, the formation of disinfection byproducts (DBPs), particularly N-nitrosodimethylamine (NDMA), has been well studied due to the carcinogenic characteristics of NDMA [276] [277] [278] [279] [280] [281] . BACs contain a quaternary nitrogen atom which can serve as a possible precursor for the formation of DBPs [282] [283] [284] . BACs are used as algaecides in swimming The biodegradation pathways for BACs in wastewater and sludge have not yet been well established. Some research papers have shown degradation mechanisms along with their byproducts and metabolites in the wastewater treatment systems during chemical oxidation processes [273] [274] [275] [276] . In water-treatment systems, the formation of disinfection byproducts (DBPs), particularly N-nitrosodimethylamine (NDMA), has been well studied due to the carcinogenic characteristics of NDMA [276] [277] [278] [279] [280] [281] . BACs contain a quaternary nitrogen atom which can serve as a possible precursor for the formation of DBPs [282] [283] [284] . BACs are used as algaecides in swimming The biodegradation pathways for BACs in wastewater and sludge have not yet been well established. Some research papers have shown degradation mechanisms along with their byproducts and metabolites in the wastewater treatment systems during chemical oxidation processes [273] [274] [275] [276] . In water-treatment systems, the formation of disinfection byproducts (DBPs), particularly N-nitrosodimethylamine (NDMA), has been well studied due to the carcinogenic characteristics of NDMA [276] [277] [278] [279] [280] [281] . BACs contain a quaternary nitrogen atom which can serve as a possible precursor for the formation of DBPs [282] [283] [284] . BACs are used as algaecides in swimming pools in combination with chlorination and UV-irradiation disinfection treatments. The cumulative process of disinfection proceeded to the formation of DBP intermediates through radical substitution reactions in the long alkyl chain of BACs. The primary products have been shown to serve as a source and precursor for the final DBPs of BACs [283] . Microbial biodegradation of BACs has been reported to produce benzyldimethylamine as the primary metabolite through the initiation of N-dealkylation at the active site of the quaternary nitrogen atom [285] . The alky intermediates in all instances were converted to carboxylic acids and the ultimate end products were obtained by the β-oxidation metabolic pathway through the citric acid cycle [285, 286] . Bacterial metabolism of BACs by sewage and soil microorganisms suggested that the terminal carbon of the alkyl moiety is oxidized to carboxylic acids, with further oxidation to end products [286] . The amine intermediates provide the final end products of carbon dioxide, water and ammonia by the metabolic mechanisms of dealkylation and deamination [287] [288] [289] . The metabolism of BACs in the human body is key for its excretion, persistence and health effects. The un-metabolized portion of BACs in the blood stream will exert systemic toxicological effects. A recent study using human hepatic cytochrome P450 enzymes (CYP) responsible for xenobiotic and drug metabolism revealed that the metabolism takes place in the long-chain alkyl end, unlike N-dealkylation cleavage at the quaternary ammonium active site [290] . CYP-mediated alkyl chain oxidation at ω and ω-1 carbon positions rendered hydroxy, dihydroxy, ketone and carboxylic acid metabolites. The current literature-reported metabolic pathways are summarized in Figure 8 . BACs with positively-charged nitrogen atoms tend to accumulate in sludge due to their strong biosorption affinity [291, 292] . BACs have been identified with a selected group of surfactants in treated and untreated sludge [293] . A summary of BACs found in sludge are presented in Table 6 . Kümmerer et al. [294] reported that the annual discharge of BACs from different European hospitals to sewage effluents was approximately 1000 kg. The emitted amount ultimately accumulated in sludge and it was found that the total sum of BAC-12 and BAC-14 in the digested sludge of three different Swedish WWTPs was greater than 1000 kg/yr, highlighting potential concerns for the use of sludge for land application [257] . Furthermore, sorption kinetic studies of municipal sludge and lime-stabilized biosolids revealed that the sorption affinity of BACs is very high, while mobility in aqueous phase is very low [295, 296] . The presence of the commonly encountered BAC-10, BAC-12, BAC-14 and BAC-16 in digested sludge provides evidence of the wide distribution of BAC homologues in sludge [261] . The total concentrations of BACs in Chinese sewage sludge were estimated in the range of 0.09-191 µg/g dry weight [256] . Sewage sludge-derived contamination has been considered to be the cause for the occurrence of BACs in soils and sediments [255, 297, 298] . The concentrations of BAC-12 and BAC-14 in the Taiwanese river Tam Sui were 2.1 and 0.8 µg/L, respectively [299] . In recent research, the average concentration of BAC-12 in wastewater influents and effluents in China was 1500 and 5.1 ng/L, respectively. The influent and effluent concentrations for BAC-14 were 586.7 and 2.5 ng/L, respectively [264] . The average concentrations of the downstream river water samples for BAC-12 and BAC-14 were 1.6 and 1.2 ng/L, respectively [264] . The biodegradation metabolites benzyl methyl amine and dodecyldimethyl amine of BACs were also identified from the returned activated sludge of a domestic WWTP in Ontario, Canada [289] . The ultimate sink of BACs is wastewater treatment systems through which they can enter into different environmental compartments [255, 292, 302] . The key transport mechanism is dissolution to the aqueous phase through cleaning and washing [263] . Cationic surfactants also contribute to the solubility and mobility of other co-existing compounds in the sewage [256] . In wastewater treatment processes, hydrophobic and electrical interactions with organic matter predominantly bind these compounds for physical separation [254, 268, 291, 296] . The longer the alkyl chain, the higher the Log K ow values and, hence, the binding affinities. The reported Log K ow values for BAC-12, BAC-14 and BAC-16 are 2.93, 3.91 and 4.89, respectively [264, 271] . The environmental fate and transformations are governed by a combination of physical, chemical and biological factors [255] . The diagenesis process also takes place in the sorbed fractions in the sludge and sediment over time [258, 302] . Biodegradation and abiotic transformations are not significant for BACs [270, 298] . Therefore, physical separation by sorption is an important contributor to the fate and environmental distribution of BACs [291, 295, 296, 303] . Benzalkonium chlorides may have adverse impacts on aquatic organisms in the environment. BAC-12, BAC-14 and BAC-16 demonstrated acute 96-h EC50 toxicity values of 0.597, 0.473 and 0.406 µmol/L, respectively, on the microalgae Chlorella vulgaris [271] . Kreuzinger et al. [304] reported EC50 values of as low as 41 µg BAC/L (72-h) for algae (Pseudokirchneriella subcapitata), 125 µg BAC/L (48-h) for rotifers (Brachionus calyciflorus) and IC50 values as low as 41 µg BAC/L (48-h) for daphnids (Daphnia magna). The potential for these compounds to have an impact will likely increase as environmental concentrations rise due to limited removals during wastewater treatment and increased use of these compounds as manufactures shift away from legacy antimicrobials. The unique physical and chemical properties of BACs allow them to be used for a variety of applications and they have become popular in a large number of industrial and domestic applications. Due to the recent outbreak of the novel coronavirus, massive hand sanitization and sterilization led to a drastic rise in the market of BACs in order to stop the spread of the disease and to protect public health [252] . The health effects due to the excessive use of these high-production-volume chemicals are of grave concern. Some potential health risks associated with the use of these products include birth defects, fertility issues, skin irritation, allergies, cancer and genotoxicity [290, [305] [306] [307] [308] . Among the QACs, benzalkonium chlorides (BACs) are one of the major classes which possess antimicrobial properties and relatively high lethal dose (LD50) values in rats (Table 7) . A recent report linked BACs in disinfectant products to developmental defects in mice, while other researchers have found that QACs can negatively impact cellular processes [309] . The purpose of the use of the disinfectants is to kill or inactivate harmful organisms. However, destroying or disturbing beneficial microorganisms in biological wastewater treatment processes cannot be avoided. Therefore, the potential risk goes beyond the intended boundary to human health and to the environment [310] . There are also increasing concerns regarding the development of antibiotic-resistant bacteria from the environmental release of BACs [288, [312] [313] [314] [315] [316] [317] [318] . BACs are in the class of halogenated organic compounds that can be persistent in the environment [298] . The presence and wide distribution of BACs in aquatic water resources, sludge, sediment and agricultural land poses multiple risks to public health and to the environment [258, 260, 265, 305, 317, [319] [320] [321] [322] [323] . Considering these risk factors and challenges, the U.S. FDA already issued a ban on one of the active quaternary ammonium ingredients, benzethonium chloride [248] . The antimicrobial ingredients of BACs may potentially be harmful, similar to banned ingredients [324] . Therefore, it is imperative that any potential risks from emerging antimicrobials are understood well-before their use becomes even more common and before these chemicals become even more widespread in the environment. Legacy antimicrobials have been a fixture of modern life for over 40 years. Two of the most common legacy antimicrobials are TCS and TCC, which have toxic, persistent and bioaccumulative characteristics. These chemicals are widespread environmental contaminants due to large production volumes, heavy usage, limited degradation during wastewater treatment and their persistence in the environment. Although aqueous phase removals of up to 98% of TCS and 99% of TCC during wastewater treatment have been reported, the primary removal mechanism is partitioning to wastewater sludge. Although some TCS and TCC removals have been observed during various forms of biological sludge treatments (typically <40% during conventional anaerobic digestion), concentrations up to 19,000 ng/g dw of TCS and 63,000 ng/g dw of TCC have been reported. Furthermore, while a number of biotic and abiotic processes have been shown to be somewhat effective in degrading these chemicals, they often generate significant quantities of metabolites and byproducts, many of which may be even more toxic and environmentally persistent than their parent antimicrobials. Recent studies are increasingly suggesting that these chemicals and their byproducts may not be as innocuous as was first thought and may have adverse short-and long-term environmental and human health impacts. Several emerging alternative antimicrobials, including various benzyl-substituted quaternary ammonium chlorides, are rapidly replacing legacy antimicrobials due to bans on active ingredients and shifting consumer preferences. There are growing concerns that many of these replacements and their byproducts present similar environmental and human health concerns as legacy antimicrobials. Few studies to date have adequately explored these alternatives so further work is urgently needed to ensure that legacy antimicrobials are not replaced by chemicals which may present similar or even possibly even greater risks. The authors declare no conflict of interest. The funders had no role in the design of the study; in the collection, analyses, or interpretation of data; in the writing of the manuscript, or in the decision to publish the results. 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