Insights into the Fate and Removal of Antibiotics in Engineered

Critical Review pubs.acs.org/est

Cite This: Environ. Sci. Technol. 2019, 53, 7234−7264

Insights into the Fate and Removal of Antibiotics in Engineered Biological Treatment Systems: A Critical Review Akashdeep Singh Oberoi,†,‡,# Yanyan Jia,†,§,# Huiqun Zhang,†,‡ Samir Kumar Khanal,∥ and Hui Lu*,†,‡

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†

School of Environmental Science and Engineering and ‡Guangdong Provincial Key Laboratory of Environmental Pollution Control and Remediation Technology, Sun Yat-sen University, Guangzhou, 510006, China § Department of Civil and Environmental Engineering, The Hong Kong University of Science and Technology, Clear Water Bay, Hong Kong ∥ Department of Molecular Biosciences and Bioengineering, University of Hawaii at Manoa, 1955 East-West Road, Honolulu, Hawaii ̅ 96822, United States S Supporting Information *

ABSTRACT: Antibiotics, the most frequently prescribed drugs of modern medicine, are extensively used for both human and veterinary applications. Antibiotics from different wastewater sources (e.g., municipal, hospitals, animal production, and pharmaceutical industries) ultimately are discharged into wastewater treatment plants. Sorption and biodegradation are the two major removal pathways of antibiotics during biological wastewater treatment processes. This review provides the fundamental insights into sorption mechanisms and biodegradation pathways of different classes of antibiotics with diverse physical−chemical attributes. Important factors affecting sorption and biodegradation behavior of antibiotics are also highlighted. Furthermore, this review also sheds light on the critical role of extracellular polymeric substances on antibiotics adsorption and their removal in engineered biological wastewater treatment systems. Despite major advancements, engineered biological wastewater treatment systems are only moderately effective (48−77%) in the removal of antibiotics. In this review, we systematically summarize the behavior and removal of different antibiotics in various biological treatment systems with discussion on their removal efficiency, removal mechanisms, critical bioreactor operating conditions affecting antibiotics removal, and recent innovative advancements. Besides, relevant background information including antibiotics classification, physical−chemical properties, and their occurrence in the environment from different sources is also briefly covered. This review aims to advance our understanding of the fate of various classes of antibiotics in engineered biological wastewater treatment systems and outlines future research directions.

1. INTRODUCTION

the largest fraction of antibiotics consumed. On the basis of the current consumption rate, global antibiotics use is projected to increase by another 15% by 2030.8 This dramatic increase in global consumption is primarily driven by rising population, rapid urbanization, and higher occurrence of infectious diseases.8 Additionally, antibiotics use for livestock production (cattle, swine and poultry) was estimated to be 118,940 t in 2013, and is expected to increase by 52% by 2030, with China being the leading consumer followed by the United States.9,10 TCs were the most commonly used class of antibiotics for

Antibiotics are critically important pharmaceuticals and are extensively used for prevention and treatment of infectious diseases in both humans and animals.1,2 Antibiotics are also widely used as a growth promoter in animals including cattle, swine, poultry, and fish.3,4 Among the varied classes, β-lactams (BLAs), fluoroquinolinones (FQs), tetracyclines (TCs), macrolides (MLs), and sulfonamides (SAs) are the most commonly prescribed antibiotics.5−7 Antibiotics consumption (expressed as defined daily dose (DDD) and data from 76 countries) increased by 65% between 2000 and 2015 from approximately 21.1 to 34.8 billion DDD. The maximum increase in consumption was observed in low- and middleincome countries including India (103%) and China (79%), where cephalosporins (37% and 24%, respectively) and broad spectrum penicillins (21% and 32%, respectively) constitute © 2019 American Chemical Society

Received: Revised: Accepted: Published: 7234

February 22, 2019 May 15, 2019 June 3, 2019 June 3, 2019 DOI: 10.1021/acs.est.9b01131 Environ. Sci. Technol. 2019, 53, 7234−7264

Critical Review

Environmental Science & Technology

parameters influencing the antibiotics removal are highlighted. Furthermore, the recent advances in engineered biological wastewater treatment systems for antibiotics removal are also discussed. Finally, the future perspectives and research gaps are outlined.

animal production both in the United States (32%) and the European Union (33%).11,12 Only a small fraction of the administered human and veterinary antibiotics are metabolized or absorbed by the body, and high percentage (50−90%) of the consumed antibiotics is excreted via urine and feces.13−15 As a result, the released antibiotics and their metabolites subsequently find their way into the environment. Additionally, the direct disposal of unused or expired antibiotics in toilets/drains and in household/hospital solid wastes also contributes to the antibiotics load in the environment.16−18 Thus, the principal sources of antibiotics in the environment include (i) sewage (treated/untreated), (ii) hospitals, (iii) livestock farms/ operations (cattle, swine, and poultry), (iv) aquaculture farms, and (v) pharmaceutical industries. Various groups of antibiotics have been frequently detected in the effluents of municipal wastewater treatment plants,2,19 secondary sludge and biosolids,20,21 surface water,22 groundwater,23 drinking water,24 and soil and sediments.25,26 Occurrence of extremely high concentrations of antibiotics has been reported in the effluent from the antibiotics production facilities (oxytetracycline (OTC): 32.0 mg/L), livestock (swine) farms (OTC: 2.1 mg/L), hospitals (ciprofloxacin (CIP): 0.9 mg/L), and in municipal wastewater (CIP: 0.25 mg/L).27−30 The highest concentration of sulfamethoxazole (SMX: 5.6 mg/L) in surface water in the vicinity of aquaculture farms (shrimp ponds) was reported.31 Residual antibiotics (mg/kg dry weight (wt)) were also detected in sewage sludge (ofloxacin (OFX): 15.1), animal (swine) manure (enrofloxacin (ENRX): 1420.7 and chlorotetracycline (CTC): 754.4), and manure-amended soils (CTC: 86.5).32−35 The presence of antibiotics and their transformation products in the environment induces the development of antibiotic-resistant bacteria (ARBs) and genes (ARGs), which may pose serious risk to both human and animal health.36 Antibiotics originating from all the sources discussed above (except aquaculture) are ultimately discharged into wastewater treatment plants (WWTPs).37,38 Several studies have shown that the conventional engineered biological treatment systems are only moderately effective in antibiotics removal, and thus high concentrations of antibiotics are subsequently being released into the environment through effluent discharge and sludge disposal.39,40 Although antibiotics are removed at varying levels during biological wastewater treatment processes, their removal pathways and mechanisms are still not clearly understood. Several laboratory41,42 and pilot-scale3,19 studies suggested that adsorption and biodegradation are the primary mechanisms for antibiotics removal from wastewater in activated sludge process. In recent years, several critical reviews examined the fate and removal of antibiotics in engineered biological treatment systems43−49 and provided information about biodegradation pathways for different sulfonamide antibiotics.50−52 However, an in-depth analysis of adsorption mechanisms, and biodegradation pathways and intermediate products for different classes of antibiotics with diverse physical−chemical properties is still lacking. Thus, the aim of this review is to provide insight into the sorption and biodegradation mechanisms of diverse antibiotics and their intermediates, and to elucidate the important factors affecting their sorption and biodegradation behaviors. Subsequently, the performance of several conventional and advanced biological treatment systems with respect to antibiotics removal is also examined. The critical operating

2. ANTIBIOTICS CHARACTERISTICS 2.1. Antibiotics Classification and Mechanisms of Action. Antibiotics are regarded as J01 drugs (antibacterial for systemic use) according to the World Health Organization (WHO) anatomic and therapeutic chemical classification.53 The mode of action of antibiotics may either be bactericidal or bacteriostatic, and is distinguished as broad or narrow spectrum based on their treatment selectivity. Depending on the types, the targets of the antibiotics may vary such as (i) cell wall and membrane, (ii) ribosomes, (iii) nucleic acids, (iv) bacterial cellular metabolism, and (v) bacterial cellular enzymes. Moreover, each group of antibiotics has different modes of action for inhibiting the bacterial growth such as (i) inhibition of cell wall synthesis, (ii) disruption of cell membrane function, (iii) inhibition of protein synthesis (both 50s and 30s ribosomal units), (iv) inhibition of nucleic acids (DNA and RNA syntheses), and (v) action as antimetabolites.54 The classification of antibiotics and their mode of action are summarized in Figure S1 in Supporting Information. 2.2. Physical−Chemical Properties of Antibiotics. Antibiotics are classified into several classes, namely, β-lactams, quinolones, tetracyclines, macrolides, and sulfonamides, which differ by their molecular structure and physical−chemical properties. Antibiotics removal through both biotic (biodegradation) and abiotic (sorption and volatilization) processes in the engineered biological treatment systems is primarily governed by their physical−chemical properties such as solubility, dissociation constant, partition coefficient, speciation, and hydrophobicity. Antibiotic compounds often contain a multitude of ionic functional groups and acid dissociation constants within a single molecule resulting in a change in physical−chemical and biological properties, such as octanol−water coefficient (log Kow), sorption behavior, photoreactivity, antibiotic activity, and toxicity with pH. Additionally, because of their low Henry’s constant values, removal of antibiotics via volatilization is often negligible.55 There are several classifications of antibiotics, and a more detailed discussion on the mode of action of antibiotics and their physical−chemical properties can be found elsewhere.56,57 This review focuses on five major classes of antibiotics (BLAs, FQs, SAs, MLs, and TCs) that are mostly used by humans and in livestock productions. The details of their physical−chemical properties are provided in Supporting Information (see pp S6−S7 and Table S1). Names and acronyms of selected target antibiotics belonging to different therapeutic classes considered in this review are also provided in Table S2 in Supporting Information. 2.3. Global Consumption of Antibiotics. There has been a perpetual increase in antibiotics consumption globally over the years in terms of both total quantity use and their types.58 There was a nearly 82%, 69%, 13%, and 8% increase in antibiotics consumption (DDD/1000 individuals/day) in China, Tunisia, India, and Greece, respectively between 2010 and 2015, whereas the antibiotics usage in the United States dropped by nearly 13% during the same period.59 The broad spectrum penicillins were the most commonly consumed class 7235

DOI: 10.1021/acs.est.9b01131 Environ. Sci. Technol. 2019, 53, 7234−7264

Critical Review


Figure 1. Potential sources and fate of antibiotics in the environment (sources−pathways−receptors (soil, surface water, and groundwater)).

productions are summarized in Supporting Information (pp S8−S10).

of antibiotics for humans with a 36% increase in their global consumption between 2000 and 2015.8 The total antibiotics usage for humans (DDD/1000 individuals/day) and consumption patterns (% of total antibiotics use) of 16 different groups of antibiotics for 10 major countries for years 2010 and 2015 are summarized in Figure S2 (Supporting Information). The antibiotics usage in animals outweighs the human consumption in many countries with an estimated global consumption of 118,940 t in 2013 and is expected to increase by 52% by 2030 due to rising global demand for livestock products (meat and dairy).10,60 The five major countries projected to share the largest consumption of antibiotics for livestock production by 2030 include China (30%), USA (10%), Brazil (8%), India (4%), and Mexico (2%).9 Total usage of all antibiotics for animal production in different geographical parts of the world is summarized in Table S3 (Supporting Information). More detailed discussions on antibiotics consumption both by humans and in animal

3. POTENTIAL SOURCES, PATHWAYS, AND OCCURRENCE OF ANTIBIOTICS IN THE ENVIRONMENT Point sources including wastewater from households, hospitals, animal (livestock and aquaculture) productions,61−63 pharmaceutical industries,64 and nonpoint sources such as agricultural and urban runoffs are the significant contributors of antibiotics in the environment.65−67 Several studies examined the occurrence and distribution of pharmaceutical compounds in different environmental matrices.1,3,68−70 There is, however, a lack of critical reviews focusing on diverse classes of antibiotics, and their occurrence in the environment from various sources. The potential sources/influxes, pathways of antibiotics in the environment, and the environmental receptors are presented in Figure 1. Concentration ranges of the most frequently detected antibiotics in various waste streams and environmental 7236


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0.0027−5528.9 0.00073−251.1 0.0018−542.5 0.0024−25.5 0.0029−361.0 0.004−68.9 0.0025−38.9 0.0006−0.5 0.0011−0.9 0.0011−2.9 0.0023−0.7 0.021−0.038 0.002−6.9 88, 114, 155−166 115.0−11000.0b 76.6−21335.0 42.0−8800.5 120.3−2810.0 9.7b−35500.0b 41.8−1908.0b 2.9−66.0b 5.1−112.3b 1.93−139.3b 1.0b−100.0b 1.8b NA 4.95−85.0 81, 82, 90, 97, 145−154 0.0053−45.6 0.0067−225.4 0.006−4.9 0.005−300.0 0.0005−3746.4 0.00013−524.4 0.00061−18.0 0.004−46.4 0.00029−20.7 0.007−0.16 NA NA 0.004−3.4 32, 116, 132−144 1.86−31000.0 25.0−420.0 0.099−4120.0 0.005−2600.0 35.8−32000.0 0.682−1000.0 0.012−1340.0 0.019 0.62 0.015−7.8 120000.0 0.027 0.0001−140.0 29, 122−131 a

NA: not available. bDigested sludge samples/biosolids.

0.094−263.1 0.02−389.2 0.029−10.0 0.006−200.0 0.031−2100.0 0.35−664.0 0.0036−50.0 0.028−153.5 0.01−800.0 0.002−4.7 NAa NA 0.002−0.6 27, 114−121 0.011−900.0 0.01−46.8 0.03−200.0 0.026−4.2 0.036−3.7 0.008−0.23 0.01−710.0 0.01−1.3 0.007−1.7 0.013−7.6 0.043−2.5 0.08−15.9 0.01−32.4 28, 61, 99−113 0.015−246.1 0.032−25.3 0.03−7.7 0.011−12.3 0.011−30.1 0.012−15.9 0.034−54.8 0.01−4.8 0.01−2.9 0.012−4.1 0.016−6.5 0.02−13.8 0.015−72.9 3, 30, 80−92 CIP NOFX OFX TET OTC CTC SMX SDZ SMZ ERY AMX CEFX TMP refs

0.0015−17.7 0.0094−2.5 0.024−6.8 0.0038−6.2 0.011−2.0 0.004−1.98 0.011−8.3 0.0011−0.56 0.0004−1.2 0.0044−2.9 0.0021−1.3 0.015−5.6 0.001−1.8 3, 30, 80−86, 91, 93−98

livestock manure and soil sewage/digested sludge (μg/ (mg/kg) kg dry wt.) pharmaceutical industries effluent (μg/L) livestock effluent (μg/L) hospital wastewater (μg/L) effluent influent antibiotics

4. REMOVAL PATHWAYS OF ANTIBIOTICS IN ENGINEERED BIOLOGICAL SYSTEMS Although WWTPs are not specifically designed and operated to remove the antibiotics, several studies highlighted the potential for attenuation and degradation of antibiotics during the biological treatment processes.97,175 The removal of antibiotics in WWTPs varies greatly from very low to nearly complete depending on their classes and types of biological treatment system employed.129 However, sorption and biodegradation are the major removal pathways of antibiotics in biological wastewater treatment systems, and removal by hydrolysis and volatilization is insignificant.38,176 In the following section, sorption and biodegradation mechanisms of various antibiotics belonging to different classes are critically discussed. 4.1. Antibiotics Removal via Adsorption in Engineered Biological Systems. Adsorption plays a primary role in the removal of antibiotics in biological wastewater treatment systems in which biological sludge serves as an important reservoir of antibiotic compounds.177,178 The adsorption process is highly complex and is affected by the physical− chemical properties of the sorbent (sludge), sorbate (antibiotics), and the operating conditions of the biological treatment systems.179−181 The extent of antibiotics sorption onto the biological sludge can be described by the solid−liquid (S/L) partitioning coefficient or sorption coefficient, Kd. The higher the Kd value, the higher is the sorption capacity and vice versa.41,182 Kd values obtained from laboratory, pilot, and fullscale studies for different classes of antibiotics in different sludge systems are summarized in Table S11 in Supporting Information. Besides, the nonlinear expressions, e.g., Freundlich and Langmuir isotherms for determining the partitioning equilibrium and adsorption capacity,183,184 are presented in Table S12 in Supporting Information. Discussions on

domestic wastewater (μg/L)

Table 1. Concentration Ranges of the Most Frequently Detected Antibiotics in Various Waste Streams and Environmental Matrices

surface water (μg/L)

groundwater (ng/ L)

matrices are summarized in Table 1. Detailed information about the potential sources contributing to antibiotics load and their occurrence in the environment is provided in Supporting Information (see pp S10−S16, Tables S4−S8 and Figures S3 and S4). The development of standardized analytical methods for multiclass antibiotic residues is critical to successfully monitor and understand the fate of antibiotics in the environment.71−73 Numerous analytical methods for simultaneous detection and quantification of multiple antibiotic compounds have been developed over the past decade. The detailed discussions on sample preparation procedures and the analytical techniques commonly used for the determination of multiclass antibiotics in wastewater and sludge samples are provided in Supporting Information (see pp S16−S20, Table S9). There is significant concern of the spread of ARBs and ARGs due to the presence of antibiotics in the environment.62,67 Several excellent reviews on ARGs and ARBs occurrence, fate, and removal in wastewater treatment systems have recently been published.74−79 In this review, our focus, however, was on the fate and removal of five major classes of antibiotics in engineered biological wastewater treatment systems, with in-depth discussion on their adsorption mechanisms and biodegradation pathways. Thus, discussion on ARGs and ARBs is not included in this review. However, occurrence of various ARGs and ARBs in wastewater treatment plants is briefly summarized in Table S10 in Supporting Information.

0.52−28200.0 1.3−35300.0 0.7−367.0 38.9−400.0 19.3−31200.0 24.0−126800.0 0.8−1820.0 1.5−17.6 1.2−128.0 5.2−143.0 NA NA 0.15−60.0 88, 115, 163, 167−174

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Figure 2. Graphical illustration showing various adsorption mechanisms (showing various functional groups involved (inner circle)) including hydrophobicity dependent (e.g., hydrophobic partioning) and independent (e.g., cation exchange, cation bridging, hydrogen bonding, and surface complexation) for diverse classes of antibiotics (outer ring) and their interactions with biological sludge.

assumptions and theories of adsorption isotherms are provided in Supporting Information (see pp S20−S22). A few recent reviews highlighted the role of sorption in the removal of antibiotics in WWTPs.46,185 These reviews, however, did not provide an in-depth analysis of adsorption mechanisms for different classes of antibiotics. Thus, there is a need to critically examine the adsorption mechanisms of different classes of antibiotics including adsorption isotherms, kinetics, and thermodynamics to better gain fundamental insights into their adsorption behavior and also to identify the key process parameters affecting their adsorption. Various adsorption mechanisms for different antibiotics and their interaction with biological sludge are illustrated in Figure 2. Extracellular polymeric substances (EPS) secreted by microbial cells are the key component responsible for adsorption of antibiotics onto biological sludge.186 Thus, the interaction between antibiotics and EPS has been also extensively discussed to better understand the fate and adsorption of antibiotics in engineered biological treatment systems. 4.1.1. Sulfonamides. SAs have been detected in the concentration ranges of 0.75 μg/kg dry wt (sulfathiazole (STZ)) to 112.0 μg/kg dry wt (SMX) in biological sludge, implying that adsorption plays a crucial role in their removal in the activated sludge process.178,187,188 For instance, 31% and 19% removal of SMX and sulfadimethoxine (SDM) (both at an initial concentration of 100 μg/L), respectively, through adsorption onto biological sludge was reported.189,190 The removal of SAs from the aqueous phase is a two-step process involving initial rapid and reversible adsorption onto the sludge followed by the perpetual removal via biodegradation.191−194

SAs have a low potential for hydrophobic partitioning since their log Kow values vary from 0.09 (sulfadiazine (SDZ)) to 1.63 (SDM) at an environmental pH of 6.0−8.0.3 Biosorption of SAs onto activated sludge is mainly governed by electrostatic interactions which are closely related to their ionization behavior.190,191 Additionally, SAs contain several moieties capable of engaging in H-bonding as solely Hacceptors (−SO2−, pyrimidine N) or as H-acceptors and Hdonors (anilinic N, sulfonamidic N) which could also facilitate adsorption.195 SAs are amphoteric molecules possessing both the amine group (−NH3+) at pKa1 of 1.6−2.6 and sulfonamide group (−SO2NH−) at pKa2 of 5.0−11.0. Hence, they can be cationic, neutral, or anionic depending on the aqueous phase pH relative to the pKa values of the compound (Figure S5a in Supporting Information).190,196 For example, if the pH value is in between pKa1= 1.85 and pKa2= 5.70 of SMX, it presents predominantly in neutral form. It would become anionic when the pH value is above 5.70 of pKa2. Since the surface charge of biological sludge is predominantly negative at a pH range of 3.0−11.0, sulfonamide antibiotics in their anionic form undergo electrostatic repulsion.190,192,193 The study with three sulfonamide antibiotics, namely, SMX, SDM and sulfamonomethoxine (SMM) showed that their removal via sorption on the biological sludge at pH 6.8 was in the order of SDM (19%) > SMM (11%) > SMX (6.5%) and consistently followed the order of their predominance in neutral form (i.e., the neutral fraction to total for SDM (24%, pKa2 of 6.3) > SMM (14%, pKa2 of 6.0) > SMX (9%, pKa2 of 5.7)).191 It is important to point out that a pH above the pKa2 of diprotic acid results in predominance of the anionic form for SAs and a 7238


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of CIP was obtained when zwitterionic species were prevalent (∼90%) at a pH of 7.4, corresponding to the isoelectric point of CIP.13 However, OFX which is mostly present in anionic form (67% at pH = 7.0) undergoes electrostatic repulsion with negatively charged biological sludge and thus is poorly adsorbed. Consequently, electrostatic interactions between the positively charged FQs and negatively charged surface of biological sludge, cation exchange of the protonated amine (pKa = 8.63 of CIP) and cation bridging via adsorbed divalent ions (especially, Ca2+ on EPS of the activated sludge) with the anionic carboxyl (pKa= 6.16 of CIP) of zwitterionic species, are considered to be the main mechanisms controlling the adsorption of FQs.3,183 Furthermore, the molecular structures of FQs also affect the adsorption by biological sludge. For example, decreased sorption of OFX is likely due to larger bulky isobutoxy group at position N1, which is in contrast to smaller cyclopropyl (CIP) and ethyl (NOFX) groups. The bulky group can cause torsion of the ring and changes the planar structure of the molecule, thereby promoting the steric hindrance of possible H-bonds involved in sorption.184 The adsorption behavior of FQs is well explained by Linear and Freundlich isotherms, and the sorption coefficients and parameters (Kd, Kf, and n) from various studies are summarized in Tables S11 and S12 in Supporting Information.38,175,184 Coefficients, Kd and Kf are influenced by temperature, and the lower temperature facilitated better adsorption of FQs, indicating the exothermic nature. The thermodynamic parameters, i.e., ΔG, ΔH, and ΔS (see Table S13 in Supporting Information for more details) revealed that the FQs adsorption onto the biological sludge is a spontaneous, exothermic, and enthalpy-driven physisorption process.38,41,175 The adsorption of FQs onto activated sludge is highly pHdependent. In the pH range of 4.0−10.0, adsorption initially increases with increasing pH, reaches a maximum level at pH 6.0−8.0, and then decreases at higher pH levels (pH of 8.0− 10.0). This decrease in adsorption at higher pH is due to unfavorable repulsive forces between the anionic carboxylate group and negatively charged sites (i.e., decreased cation bridging), and the competition between the carboxylate moiety and the aqueous hydroxyl ions for surface bound exchangeable metal ions (i.e., decreased surface complexation).175,183,184,201 Antibiotics adsorption onto biological sludge occurs through both hydrophobic partitioning and hydrophobicity-independent mechanisms (e.g., electrostatic interactions), which are closely related to the sludge properties.38 Sludge characteristics depend on the types of treatment system (aerobic and anaerobic).202 There are contrasting reports in the literature with regard to antibiotics adsorption on aerobic and anaerobic sludges. For instance, some studies observed higher sorption of FQs on aerobic sludge, whereas other studies reported higher Kd values in anaerobic sludge systems.38,150,175 This is mainly attributed to variation in sludge property (i.e., organic content), biomass concentration (i.e., MLSS), and changes in the EPS content and composition in aerobic and anaerobic sludges.38,203,204 The aerobic and anaerobic conditions have been shown to influence the production of EPS, the key component responsible for adsorption of antibiotics onto biological sludge as highlighted in the next section. Some studies reported that more EPS are produced in aerobic sludge than anaerobic sludge systems,203,205 whereas Wei et al.206 found more EPS production in anaerobic sludge than aerobic sludge, especially the protein content. Additionally, differences

higher charge repulsion and lower adsorption efficiency for SMX as compared to SDM.190 The adsorption behavior of SAs at equilibrium is well described by Linear and Freundlich isotherms with n value approximating unity.181,190,192,193,197 The linear partition coefficient, Kd, and Freundlich sorption parameters, K f (sorption affinity) and 1/n (sorption intensity), are summarized in Tables S11 and S12 in Supporting Information. The thermodynamic parameters including Gibbs free energy change (ΔG), enthalpy change (ΔH), and entropy change (ΔS) are extensively used to elucidate the adsorption behavior, and the SAs adsorption by activated sludge is generally a spontaneous, exothermic, and physisorption process (see Table S13 in Supporting Information for important thermodynamic parameters).192,197 Various operating/environmental factors, such as temperature,192,197 pH,192 solids retention time (SRT),178,198 redox conditions,198 total organic carbon (TOC),178 and mixed liquor suspended solids (MLSS) concentration192 affect the sorption behavior of SAs. The ionic speciation of SAs is pHdriven, and therefore pH plays an important role in their sorption onto biological sludge. Sorption of sulfamethazine (SMZ) showed a decreasing trend with an increasing pH from 5.0 to 11.0 mainly due to electrostatic repulsion.192 SRT is also an important factor influencing SAs sorption onto biological sludge.178,192 A longer SRT generally results in higher biomass concentration and consequently an increased in the total amount of adsorbed antibiotics.192 Adsorption of SAs decreased with increasing temperature, indicating that adsorption onto sludge is an exothermic process.192 Significant correlation between TOC and total concentration of antibiotics in biological sludge was observed, thereby attributing to the potential of antibiotics to bind with organic matter in biological sludge.178 4.1.2. Fluoroquinolones. FQs such as CIP, ENRX, OFX, and norfloxacin (NOFX) are frequently detected in biological sludge at relatively high concentrations ((mg/kg dry wt) OFX (15.1), CIP (3.9), NOFX (7.5), ENRX (0.09), and lomefloxacin (LOME) (1.1)).19,34,97,199 Adsorption is the primary removal pathway of FQs during biological wastewater treatment since the biodegradation rate of FQs in the activated sludge process is substantially low (only 6.3 μg/g-suspended solids (SS)/d for CIP).38,176 Adsorption and biodegradation accounted for 50−91% and 9−22% of overall removal of different FQs (CIP, ENRX, OFX, NOFX and LOME), respectively, in the activated sludge process.38,150 Despite their highly hydrophilic characteristics (log Kow ≪ 2.5), FQs (e.g., CIP, ENRX, and OFX) have a high adsorption potential onto biological sludge (Kd > 500 L/kg MLSS, see Table S11 in Supporting Information), implying that the adsorption of FQs could not simply be attributed to hydrophobic interactions.41 FQs are amphoteric molecules with two ionizable functional groups (i.e., acid 3-carboxyl group (pKa1 = 5.9−6.3) and the basic N-4 in the piperazine substituent (pKa2 ≈ 8.0)),200 which form cationic (pH of 4.0− 5.0), zwitterionic (pH of 7.0−8.0), and anionic (pH > 9.0) species in aqueous phase depending on the pH (Figure S5b in Supporting Information).191 The contribution of different species of FQs (e.g., LOME, ENRX, NOFX, CIP, and OFX) to the overall adsorption (74−98%) showed that zwitterions are the major contributors under typical pH conditions (pH of 6.5−7.5) in biological wastewater treatment systems.38,41,182,183 For instance, the highest adsorption capacity 7239


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process with the heterogeneous nature of adsorption sites.209,213,214,216,219 Kd values for different TC antibiotics in various sludge systems are summarized in Table S11 in Supporting Information. The higher Kd values indicate the higher adsorption affinity of TCs. The Freundlich coefficient and parameter (Kf and n) are summarized in Table S12 in Supporting Information. The thermodynamic analysis of TCs adsorption on biological sludge (Table S13 in Supporting Information) shows negative ΔG, ΔH, and ΔS values, manifesting that the TCs adsorption onto biological sludge is a spontaneous, exothermic, and enthalpy-driven process.197,216,218 The adsorption of TCs onto biological sludge is highly pHdependent process with a significant decrease in adsorption capacity with an increase in pH from 3.3 to 8.4 because of the speciation of TCs under different pH conditions. For instance, TET predominantly exists in anionic form at pH > 7.7, thereby causing electrostatic repulsion with the negatively charged surface of biological sludge and resulting in decreased adsorption.213,214,216,218 The presence of different ionic species (Na+, Ca2+, Mg2+) in wastewater reduced the adsorption of TCs (TET and OTC) onto activated sludge mainly due to the formation of strong complexes between TCs and cationic ions.213,214,220 Additionally, the competition between the positively charged quaternary ammonium functional group of TET and divalent cations (Ca2+ and Mg2+) for the cation exchange sites of adsorbent surface (e.g., carboxyl groups on EPS of biological sludge) could also decrease the TET adsorption.213,221,222 Interestingly, OTC adsorption enhanced in the presence of Cu2+, which was attributed to its ability to act as a bridging ion between carboxylic and amide (II) moieties of biological sludge and OTC, resulting in the formation of ternary complexes.214 These findings apparently show that the surface complexation through metal bridging also play an important role in the adsorption of TCs. Furthermore, overall Kd values (Table S11 in Supporting Information) reveal that the sludge from aerobic conditions facilitated better adsorption of TCs (doxycycline (DOXC), OTC, and TET) than the sludge from anoxic conditions. This is mainly attributed to different physical−chemical properties of sludge (e.g., EPS concentration and composition contributing different functional groups) and organic matter content.216,223 Compared to suspended activated sludge, aerobic granular sludge showed a better biosorption capacity for TCs, which was attributed to the presence of a large number of various functional groups, such as amine, carboxylic, aldehyde, hydroxyl, sulhydryl, and amide on the surface of granules.219 For instance, amide groups in peptide bonds of proteins facilitated OTC biosorption onto the aerobic granules surface.215 4.1.4. Macrolides. High concentrations (μg/kg dry wt) of azithromycin (AZI) (6027.2), roxithromycin (ROX) (110.0), clarithromycin (CLAR) (49.5), and erythromycin (ERY) (43.4) were reported in biological sludge collected from municipal WWTPs.19,97,150 Adsorption is the major removal mechanism for macrolide antibiotics in biological wastewater treatment systems.19,176,216 Hydrophobic interactions were predominantly responsible for MLs (ERY, CLAR, and AZI) sorption onto the activated sludge due to their relatively high octanol−water coefficient (log Kow of 0.73−3.09) as shown in Figure 2.3,187 Additionally, through the protonation of the tertiary amino group (pKa > 8.7), MLs (e.g., ERY, AZI, and CLAR) exist predominantly in cationic form (71−99%) under

in the chemical structure of the EPS extracted from aerobic and anaerobic sludges were also observed.205 This variation in EPS characteristics can influence the availability of adsorption sites and binding strength between antibiotics and sludge. In addition to EPS, antibiotics adsorption can also be influenced by the organic content of sludge (i.e., volatile suspended solids (VSS)/total suspended solids (TSS) ratio). Generally, anaerobic sludge tends to have a slightly lower VSS/TSS ratio than the aerobic sludge due to long SRT and accumulation of nonvolatile materials.207,208 Positive correlation between TOC and concentration of antibiotics in sewage sludge was observed, thereby attributing to a higher potential of antibiotics binding with organic matter.178 Thus, the differences in the concentration and characteristics of organic matter content between aerobic and anaerobic sludges are the key factors influencing antibiotics adsorption. Furthermore, suspended sludge was found to have a higher biosorption capacity than the granular sludge which was attributed to higher EPS content, especially proteins.184 The presence of ion species, such as Ca2+ and Mg2+, negatively affected FQs (NOFX, OFX, and CIP) adsorption onto activated sludge mainly due to the formation of stable complexes between divalent cations and FQ molecules.176 4.1.3. Tetracyclines. TCs have been detected in high concentration in biological sludge. For example (mg/kg dry wt), 3.8 of OTC, 1.9 of CTC, and 0.4 of tetracycline (TET) in biological sludge from municipal WWTPs, 34.5−86.4 of OTC and 4.1−28.3 of CTC in biological sludge from livestock WWTP were reported.150,151,209 The removal of TCs from aqueous phase involves initial rapid adsorption onto biological sludge followed by biodegradation, especially in nitrifying sludge systems and is attributed to long SRTs (≥18 days).210−212 Similar to SAs, TCs possess high water solubility and low log Kow; therefore, non-hydrophobic interactions govern the transport and fate of TCs in biological wastewater treatment systems and their overall removal.176,197,213−216 TCs possess multiple ionizable functional groups, i.e., tricarbonyl amide (C-1/C-2/C-3), phenolic diketone (C-10/ C-11/C-12), and dimethylamine (C-4) groups, which correspond to three acid dissociation constants with pKa values of 3.3, 7.7, and 9.7, respectively.213,217 Thus, TCs exist as cationic, zwitterionic, and anionic species under acidic, moderately acidic to neutral and alkaline conditions, respectively (Figure S5c in Supporting Information). The adsorption affinity of different ionic species onto biological sludge follows the order of cationic ≫ zwitterionic > anionic.213,216 Under typical pH conditions (6.5−7.5) in biological wastewater treatment systems, adsorption of TCs primarily occurs via electrostatic interactions between the zwitterionic species and negatively charged surface of biological sludge.179,214,218 The contribution of zwitterionic species to the overall adsorption is often over 90% in the pH range of 6.0−7.0 in an activated sludge process.213,216 The direct surface complexation between OTC and different organic functional groups of biological sludge, such as carboxylic acids (−COOH), aldehyde (−COH), hydroxyl (−CHOH), and amine (−NH2), plays an important role in the adsorption process (Figure 2).215,216 The adsorption of TCs decreased at a higher temperature, suggesting an exothermic reaction, which is attributed to decreased Van der Waals forces.197,215,218,219 Both Linear and Freundlich isotherms describe well the adsorption behavior of TCs onto biological sludge, suggesting a multilayer adsorption 7240


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route for CEFX, while adsorption was the major removal mechanism for AMX and AMPI.176,230 Electrostatic interactions played a dominant role in the sorption of penicillin antibiotics onto the biological sludge. For instance, AMX is an amphoteric molecule with three ionizable functional groups (i.e., carboxyl, amine, and hydroxyl with their respective pKa values of 2.6, 7.4, and 9.6).231,232 Consequently, within the pH range of 3.0−7.0 (i.e., pKa1 < pH < pKa2, pKa3), amine functional groups are present as −NH3+ ,and carboxyl functional groups are present as − COO− (Figure S5d in Supporting Information). Therefore, under typical pH conditions (6.5−7.5) in WWTPs, electrostatic interaction between zwitterions and negatively charged surface of biological sludge likely plays a dominant role in the sorption of AMX. Adsorption capacity of activated sludge for penicillin G (PEN G) increased with temperature from 25 to 35 °C, indicating the endothermic nature of sorption in this temperature range; however, the adsorption capacity decreased significantly with a further increase in temperature due to the decreased surface activity and rapid decomposition of PEN G at higher temperature. The endothermic nature of adsorption was further confirmed by the positive ΔH. The negative ΔG indicates the spontaneous nature of the adsorption process.233 The thermodynamic parameters are summarized in Table S13 in Supporting Information. Role of EPS on Antibiotics Adsorption. EPS are mixture of high molecular weight polymers produced by microorganisms.234−236 EPS have complex and diverse composition in terms of proteins (PN), polysaccharides (PS), nucleic acids, lipids, and humic substances, which contribute to different functional groups, such as carboxyl, amine and hydroxyl, and hydrophobic regions, thereby providing numerous binding sites for the adsorption of diverse organic micropollutants.234,235 The following section critically discusses the role of EPS in antibiotics removal by sorption and underlying mechanisms of the interaction between EPS and different antibiotics. EPS production in biological sludge is greatly influenced by the concentration of antibiotics, thereby elucidating the role of EPS in microbial responses to antibiotics exposure. Microbes tend to secrete more EPS as an adaptive mechanism in response to the toxicity of the antibiotics.203,237−239 The protein-to-polysaccharide (PN/PS) ratio increases with an increase in antibiotics concentration (Table S14 in Supporting Information). EPS also have highly desirable sorption characteristics for micropollutants, such as antibiotics.203,239−241 Antibiotics could be effectively adsorbed on EPS, which accounted for up to 15%, 88%, and 13% of total sulfamethizole (SMT), TET, and NOFX (at an initial concentration of 1000 μg/L) adsorbed by biofilm, respectively, and up to 36% and 35% of total CIP (at initial concentration of 1000 μg/L) and SMZ (at initial concentration of 500 μg/L) adsorbed by aerobic sludge, respectively.186,203,239 EPS extracted from various biological sludge systems (such as aerobic, anaerobic (methanogenic), aerobic granular, and sulfate-reducing bacteria (SRB)) show interaction with different antibiotics (e.g., SMZ, CIP, TET, and ERY) to form a stable EPS−antibiotics complex through hydrophobic interaction, hydrogen bonding, or electrostatic interaction.186,203,239−241 The mechanism of interaction and the binding process between EPS and the antibiotics is commonly examined using fluorescence quenching assays

typical pH conditions (6.0−8.0) in biological wastewater treatment systems. Consequently, these cations tend to get sorbed easily on the negatively charged activated sludge via electrostatic interactions.19,180,216 The role of electrostatic interactions in sorption of MLs was also confirmed by Wunder et al.224 Antibiotic speciation and molecular size are the two important factors affecting the interactions between antibiotics and biofilm, and the extent of sorption was in the order of CIP ≫ ERY > SMX.224 SMX is predominantly neutral to negatively charged at environmental pH (6.5−7.5),190 while CIP and ERY are predominantly positively charged,199,216 thereby resulting in enhanced sorption onto negatively charged biological sludge. Furthermore, the reduced extent of sorption of ERY relative to CIP was likely due to the larger molecular size of ERY that resulted in a mass transfer (i.e., diffusion) limitation in the biofilm.224 The biomass physical conformations (i.e., suspended sludge and biofilm) are expected to have an impact on the fate of antibiotics, as it affects mass transfer between the target compounds and the microorganisms, and the release of metabolites into aqueous phase from biomass aggregates.45 The sorption coefficients of three MLs (CLAR, ROX, and ERY) in biofilm were comparable or higher than that in activated sludge, which was attributed to higher accessible surface area associated with the biofilms porous structure.225 Moreover, positive correlation between antibiotics adsorption (CLAR, ERY, and ROX) and biofilm thickness was reported.225 For example, Kd values (L/g) for CLAR were 0.42, 0.41, and 11.20 corresponding to biofilm thickness of 50, 200, and 500 μm, respectively. This was possibly due to lower biomass density and significantly higher porosity and accessible surface area in the thickest biofilm.225 In the case of the biofilm, sorption behavior is best described by the intraparticle diffusion model comprising a two-stage process: initial quick sorption of organic pollutants onto the surface of the biofilm followed by a slow intramolecular diffusion toward the inner layers.226,227 The time required to achieve partitioning equilibrium in suspended activated sludge system is relatively faster compared to biofilms system, where molecular diffusion can significantly influence the mass transfer and partitioning kinetics.225,228 However, little is known about the sorption of antibiotics on biofilms. The sorption of MLs (AZI) onto the biological sludge was a reversible, spontaneous, exothermic, and enthalpy-driven process governed by the surface interactions.216 The sorption behavior of MLs conformed well with both Henry and Freundlich models.216 Kd values for different MLs in various sludge systems are presented in Table S11 in Supporting Information. 4.1.5. β-Lactams. β-Lactam antibiotics are often detected in biological sludge from WWTPs at a very low concentration. For example, amoxicillin (AMX) and ampicillin (AMPI) concentrations of 1.0 and 14.8 μg/kg dry wt, respectively, were reported in biological sludge collected from municipal WWTPs.149 The β-lactam ring is highly unstable and easily cleaved by β-lactamases, a group of enzymes widely excreted by bacteria.57,229,230 Notably, studies on transformation and fate of β-lactams in biological wastewater treatment systems showed that the removal routes for major antibiotics in this class, including AMX, AMPI (penicillin), and cephalexin (CEFX) (cephalosporins), varied greatly, possibly due to their different chemical structures, such as variable side chains. Biodegradation was found to be the predominant removal 7241


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Environmental Science & Technology including static and dynamic quenching,186,203,241 and the biomolecular quenching rate constant (Kq) is determined to show the interaction process. The quenching of fluorescence intensities of EPS fluorophore (proteins and humics) with the increasing antibiotics concentration may likely be attributed to the formation of stable EPS−antibiotics complexes (static quenching) by changing the conformational structure of EPS.186,203 Additionally, the Kq values for both proteins and humic-like substances are far greater than 2.0 × 1010 L/mol/s, the maximum diffusion collision quenching rate constant of various quenchers with the biological macromolecules.242,243 Protein-like substances such as tryptophan and tyrosine are the dominant components of the EPS that interact with the antibiotics thereby facilitating their sorption onto biological sludge through hydrophobic interactions (Table S15 in Supporting Information).186,237,239,240 The humic-like substances of EPS are also involved in sorption of antibiotics, including CIP, ERY, and STZ.237,239,241 The binding constant (Kb) which represents the affinity between EPS components and antibiotics,241,243 indicates that the protein-like substances have a higher binding affinity to antibiotics (Kb = 1.04 × 103 to 5.04 × 105) than those of humic-like substances (Kb = 3.09 × 102 to 9.33 × 102) (see Table S15 in Supporting Information for Kb values). Therefore, proteins as the dominant component in EPS have a stronger interaction with antibiotics over humiclike substances. The hydrophobicity of EPS is usually characterized by the mass ratio of PN/PS in EPS, and the higher PN/PS ratio indicates the stronger hydrophobicity and more availability of adsorption sites.234,244 Compared with the EPS extracted from aerobic and anaerobic (methanogenic) sludge, the EPS from SRB sludge have a higher PN/PS ratio, indicating stronger hydrophobicity, and higher availability of adsorption sites (binding sites, n) and binding strength (binding constant, Kb) for CIP adsorption (Table S15 in Supporting Information).203 Moreover, carboxyl, amine, phenolic, and hydroxyl groups of EPS were the dominant functional groups involved in antibiotics (CIP and TET) adsorption.203,240 Thus, both hydrophobic partitioning and hydrophobicity-independent mechanisms (e.g., electrostatic interaction, cation exchange, cation bridging and surface complexation) are involved in the binding of antibiotics to EPS.186,203,240 The Freundlich isotherm describes well the sorption behavior of EPS for antibiotics (CIP, TET, and STZ),203,239,245 implying the nonlinear and the heterogeneous nature of antibiotics adsorption onto EPS. This is primarily due to the complex structure of EPS and the heterogeneous distribution of the adsorption sites in EPS as reflected by Kf and 1/n values (Table S12 in Supporting Information). Antibiotics adsorption is mostly an exothermic process, and it decreases with increasing temperature, such as the adsorption of TET and SMZ by EPS in activated sludge, and the driving force is dominated by hydrophobicity and electrostatic force (hydrogen bond in the binding interaction).186,240 The effectiveness of antibiotics sorption by EPS depends on several factors such as pH, temperature, ionic strength, types and concentration of the antibiotics, and the types of sludge system.203,245 The protonated or deprotonated states of different functional groups of both EPS and antibiotics vary considerably with pH, thereby significantly influencing the adsorption capacity of EPS.245 The EPS from various sludge systems have multiple components (such as proteins and

polysaccharides, and their ratio (PN/PS)), which may result in varied sorption capacity and affinity.203 Understanding the role of EPS from different biological sludge systems on sorption of antibiotics will help better understand the migration and fate of antibiotics in engineered biological treatment systems. However, information on the interaction mechanisms between EPS from various biological sludge systems and antibiotics is still very limited. 4.2. Biodegradation Mechanisms and Pathways. As discussed in the preceding section, sorption onto biological sludge in biological wastewater treatment systems plays a key role in antibiotics removal from the aqueous phase. However, sorption is a phase transfer phenomenon, and the risk of antibiotics release to the environment still exists. For some antibiotics of sulfonamides class (e.g., SMX and SDZ) and trimethoprim (TMP), sorption onto biological sludge plays a minor role in their removal in engineered biological treatment systems,190,191 and therefore, biodegradation is the principal pathway for their removal. Biodegradation is a process that involves the breakdown of complex organic compounds including antibiotics either through biotransformation, resulting in formation of different metabolic intermediates (i.e., dead end and/or transitory intermediates)246−250 or through complete mineralization to CO2 and H2O247,251,252 by pure or mixed microbial culture. The terms biotransformation and biodegradation are interchangeably used in the literature. Different intermediates could be formed either by breakdown of the parent antibiotic compounds246,247,249,253 or by hydroxylation, acetylation of the amino group in SMX,250,254−256 and loss of N-methyl group by demethylation of the dimethyl amino group at C4 position of TET molecule257 without breakdown of the parent compound. Microorganisms have the ability to degrade antibiotics utilizing them as a sole carbon and energy source and/or via cometabolism.52,249,254 Several recent studies focused on possible biodegradation mechanisms (including intermediates, pathways, functional microorganisms, catabolic enzymes, and genes) of antibiotics in different microbial sludge systems, which are important to elucidate the fate of antibiotics in the environment.246,258−260 The following section examines the biodegradation mechanisms of the selected target classes of antibiotics. Moreover, the biotransformation and mineralization by pure and mixed cultures under different redox conditions, intermediates, pathways, catabolic enzymes, and genes involved are also discussed. 4.2.1. Sulfonamides. SAs (SMX and SDZ) are widely used for the treatment of bacterial infections in urinary tract, bronchitis, and prostatitis both in humans and animals.261,262 SMX is one of the most widely and frequently detected sulfonamide antibiotics in the environment with concentrations ranging from ng/L to mg/L.263−265 SMX is poorly adsorbed on biological sludge as discussed earlier. However, biotransformation and mineralization have been observed with both pure and mixed cultures in different redox (aerobic, anoxic, and anaerobic (sulfate-reducing, methanogenic, and iron-reducing)) conditions.251−253,255,266−268 Several pure bacterial strains including Microbacterium sp. strain BR1,246,269 Achromobacter denitrificans PR1,270 Pseudomonas psychrophila HA-4,271 and Acinetobacter sp.268 have been isolated and characterized for their ability to degrade SMX as a sole carbon and energy source under aerobic conditions. Recently, two iron-reducing Shewanella strains, i.e., Shewanella oneidensis MR-1 and Shewanella sp. MR-4, have been isolated 7242


Critical Review


Figure 3. Biotransformation products and pathways of different antibiotics belonging to diverse classes (a) SMX biotransformation by sad genes, (b) CIP biotransformation pathways in aerobic, anoxic, and anaerobic biological sludge systems and by the N-acetylation enzyme, (c) TCs biotransformation by tetX genes, (d) ERY biotransformation catalyzed by Mph, Mgt, and Ere genes, and (e) AMX biotransformation through hydrolysis of β-lactam ring (FAD = flavin adenine dinucleotide, FADH2 = flavin adenine dinucleotide H reduced form, NADP+ = nicotinamide adenine dinucleotide phosphate oxidized form, NADPH + H+ = nicotinamide adenine dinucleotide phosphate reduced form).

that demonstrated high SMX degrading potential with maximum removal of 60% and 64%, respectively, at an initial concentration of 10 mg/L.267

In an aerobic process (including pure and mixed cultures), SMX is commonly biotransformed to 3-amino-5-methylisoxazole (3A5M) (Figure 3a (i)).246,260,267,270 The inter7243


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tion.278,279 However, some studies reported CIP biotransformation products and pathways in different biological sludge systems (e.g., aerobic, anoxic, and anaerobic conditions) (Figure 3b and Figure S8 in Supporting Information).38,41,280−282 CIP degradation initiated at the piperazine ring in all the biological sludge systems (aerobic, anoxic, and anaerobic) but with the formation of different intermediates of low toxicity.258,280,282 In an aerobic condition, the piperazine ring of CIP got oxidized thereby forming carboxyl groups (Figure 3b (i)), while the quinolone part of the CIP remained intact.280 In an anoxic (denitrifying) condition, CIP degraded via the addition of the nitroso group and three oxygen atoms in the piperazine ring (Figure 3b (ii) and (iii)).281 In an anaerobic sulfate-reducing condition, CIP degradation began at the piperazine ring with the loss of C2H2 via desethylation (Figure 3b (iv)) and subsequently degraded via the loss of a C2H5N fragment (Figure 3b (v)). Moreover, additional three intermediates (Figure 3b (vi), (vii), and (viii)) were formed via hydroxylation in different positions of CIP.282 However, studies on CIP biotransformation products and pathways in anaerobic methanogenic sludge system are still lacking. Recently, CIP biotransformation was found to take place via N-acetylation catalyzed by acetylation-CIP enzyme encoded by the aac(6′)-Ib-cr gene (Figure 3b and Figure S9 in Supporting Information).258 Additionally, CIP could be degraded via sequential hydroxylation catalyzed by basidiomycetes, brown rot fungus Gloeophyllum striatum, resulting in formation of several different intermediates by a hydroxyl radical based degradation mechanism283 or via desethylene-N-acetylation and/or N-formylation reactions catalyzed by ascomycetous soft rot fungus Xylaria longipes.284 These biotransformation activities are catalyzed by extracellular lignolytic enzymes (i.e., laccase, lignin peroxidase, and manganese peroxidase) and intracellular enzymes (i.e., belonging to cytochrome P450 family). 4.2.3. Tetracyclines. TCs (TET, OTC, and CTC) are broad spectrum antibiotics used in livestock production. TCs are poorly biodegradable due to their complex chemical structures; therefore, several studies explored chemical processes (i.e., photochemical and electrochemical technologies) for their degradation.285 However, very little is known about their microbial degradation. TCs could be transformed to C11a-hydroxy-tetracyclines catalyzed by a flavin monooxygenase encoded by tetX genes in microbes (Figure 3c and Figure S10 in Supporting Information).286 The two additional orthologues of the original tetX gene (i.e., tetX1 and tetX2) have been identified encoding for enzymes involved in TCs degradation by aerobic and anaerobic bacteria.287−289 More recently, a novel bacterial strain, Stenotrophomonas maltophilia DT1, capable of degrading TET was isolated from TET contaminated soil (Figure S11 in Supporting Information for intermediates).257 Genome analysis revealed the presence of tetX1, a gene encoding flavin adenine dinucleotide (FAD) binding monooxygenase, and eight peroxidase genes, in Stenotrophomonas maltophilia strain.290 On the basis of a molecular mechanism of TET biotransformation by the S. maltophilia strain, the nodulation protein efflux pump transported TET outside cells, and hypoxanthine-guanine phosphoribosyl-transferase facilitated the activation of the ribosomal protection proteins. Finally, TET biotransformation was catalyzed by enzymes superoxide dismutase and peroxiredoxin.290

mediate 3A5M is formed due to the release of 4-iminoquinone and sulfur dioxide simultaneously from the parent compound (SMX) as elucidated in eq i): C10H11N3O3S(SMX) + OH− → C6H5NO(4‐iminoquinone) + SO2 + C4 H5N2O−(3A5M−) + 2H+ + 2e−.

(i)

This ipso-hydroxylation reaction is catalyzed by monooxygenase encoded by the sadA gene, which allows sulfonamide functional group to separate from the parent compound thereby lowering the potential impact of intermediates on the environment.252,272 In Microbacterium sp. strain BR1 and other actinomycetes, a flavin-dependent monooxygenase encoded by sadA gene and a flavin reductase encoded by sadC gene are responsible for the initial breakdown of sulfonamide molecules, resulting in the release of 4-aminophenol. The latter is further transformed into 1,2,4-trihydroxybenzene by monooxygenase encoded by sadB gene and flavin reductase encoded by sadC gene prior to mineralization (Figure 3a (ii) and Figure S6 in Supporting Information).246,252 These studies indicated that Microbacterium sp. strain BR1 was able to utilize the aniline moiety of SAs for growth and was capable of mineralizing SMX. In addition, SMX biotransformation in aerobic condition could be catalyzed by other functional enzymes (such as peroxidases, dioxygenases, and cytochrome P450 enzymes) that are involved in the aromatic ipso-subsitutions,246,255,273−275 and more intermediates are formed via different pathways (Figure S7 in Supporting Information).255,276,277 SMX could be biotransformed into N4-acetylSMX, 4-nitro-SMX, and desamino-SMX (Figure S7H−J) by enriched ammonia-oxidizing bacteria (AOB), indicating different intermediates and pathways under nitrifying conditions.255 These findings apparently indicate the preferential biodegradation of certain antibiotics under a particular redox condition by specific group of bacterial population. However, further studies are needed to gain insights into biodegradation pathways and intermediates of different antibiotics by AOBs. Anaerobic, especially the SRB sludge system showed different intermediates and pathways of SMX degradation compared to the aerobic sludge system.253 SMX degradation begins with the isoxazole ring cleavage through hydrogenation to form an unstable radical anion (SMX+) leading to the formation of different end-products (Figure S7 in Supporting Information).253 Meanwhile, 3-hydroxylamine-amino-5-carboxyl (Figure S7U in Supporting Information) formed via aromatic ipso-hydrogenation was reported for the first time in the SRB sludge system.253 These findings apparently indicate that SMX intermediates and pathways are significantly different in different biological sludge systems that are mediated by different microbial groups. Though an anaerobic process (methanogenic) is widely used for sewage sludge digestion, and livestock manure and wastewater treatment, there are very limited studies on intermediates and pathways of SMX biodegradation. Thus, more studies are needed to investigate SMX biodegradation mechanisms, pathways, and intermediates under anaerobic methanogenic conditions. 4.2.2. Fluoroquinolones. CIP is the most widely prescribed FQ antibiotic for humans and animals, and is frequently detected in the environment. The rapid removal of CIP in biological treatment systems is mainly via sorption onto biological sludge and is highly recalcitrant to biodegrada7244


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Table 2. Comparative Summary of Different Engineered Biological Wastewater Treatment Systems for Antibiotics Removal

about their biodegradability, possible biotransformation products, and pathways in biological treatment systems. A few studies reported that ERY could be degraded by pure and mixed cultures under aerobic and anaerobic conditions.296−298 Recently, a pure bacterial strain Ochrobactrum sp. isolated from ERY contaminated soil demonstrated high degradation potential with maximum removal of 97% at an initial concentration of 100 mg/L, and the biotransformation products indicated the removal of both the sugar moieties from macrolactone ring structure.298 AZI and ROX were effectively biotransformed via phosphorylation by activated sludge in membrane bioreactor system.299,300 In addition, different intermediates of AZI, ERY, and CLAR were reported to be formed via the removal of one or both sugar units or by N-oxidation, N-demethylation, and phosphorylation of the desosamine sugar moiety by an AZI-enriched activated sludge culture (Figure S13 in Supporting Information), which could be due to the involvement of different microbial enzymes.301 MLs biotransformation could be catalyzed by some specific inactivated enzymes, such as phosphotransferases encoded by the Mph gene family, and glycosyltransferases encoded by the Mgt gene family (Figure 3d and Figure S14 in Supporting Information for intermediates).302 Moreover, MLs could also be degraded via hydrolysis resulting in the cleavage of a lactone

In addition, several studies focused on TCs biotransformation using fungal species.259,291−293 The ligninolytic enzymes (e.g., laccase, and lignin and manganese peroxidases) excreted by fungi could catalyze TCs (such as TET, OTC, and CTC) biotransformation.259,291 For example, OTC was transformed to a less toxic intermediate, 2-acetyl-2-decarboxamido-oxytetracycline (Figure S12A in Supporting Information), catalyzed by laccase obtained from Pleurotus ostreatus mycelium.259 Besides, CTC was transformed into three nontoxic intermediates (Figure S12B in Supporting Information) which was catalyzed by fungal-derived laccase.293 Trichosporon mycotoxinivorans XPY-10, isolated from the antibiotic producing pharmaceutical industry wastewater, was capable of degrading TET. TET was degraded via dehydroxylation, demethylation, and deamination reactions, resulting in formation of five intermediates (Figure S12C in Supporting Information).292 However, detailed studies on biodegradation mechanisms (including intermediates and pathways) of TCs in different biological sludge systems are still lacking. 4.2.4. Macrolides. ERY along with three other MLs (CLAR, ROX, and AZI) are widely used as human medicine and in livestock production.294 Studies have reported that these MLs could be easily transformed into a large number of intermediates via photolysis.295 However, very little is known 7245


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antibiotics based on aqueous and sludge phase concentration is a more realistic approach to determine the overall removal of antibiotics and to better understand the fate and pathways of antibiotics removal in biological wastewater treatment systems. Furthermore, few studies stress the need of combined quantification of both the parent compound and the conjugate metabolites to determine the total removal efficiency, as deconjugation is critical for a number of antibiotics (e.g., SAs).316 In addition, various mathematical models have been developed to predict the fate and removal of antibiotics during biological wastewater treatment processes.317−319 Especially, activated sludge model for xenobiotics (ASM-X) has been previously developed to predict the fate of antibiotics (i.e., SMX, CIP, and TET).183,317 More recently, the ASM-X framework has been used to identify the influence of different factors on antibiotics removal during wastewater treatment.316 Polesel et al.316 showed that the variability in observed removal efficiencies could be associated with retransformation of conjugated metabolites back to parent antibiotics, SRT, adsorption on influent and effluent solids, influent−effluent dynamics and sampling strategy. More details about the models can be found elsewhere.316,317,320 5.1. Conventional Biological Treatment Systems. 5.1.1. Activated Sludge Process (ASP). ASP is the most widely adopted system for biological wastewater treatment.46,179 Though widely used, ASP is not specifically designed for the removal of antibiotics. 37,321 Several laboratory- and field-scale studies showed that the antibiotics are only partially removed in conventional WWTPs with removal efficiencies ranging from negative to over 99%.3,19,95,322 For instance, in full-scale ASP, antibiotics removal from the aqueous phase varied significantly from −1% (AZI: 0.13 μg/L), 25% (NOFX: 0.02 μg/L), 52% (SMX: 0.44 μg/L), 66% (ERY: 0.045 μg/L), to 71% (CIP: 2.2 μg/ L).322 Biosorption was the major removal pathway for FQs and MLs, whereas biodegradation was the major removal route for SAs.19,323 Sequencing batch reactor (SBR) has also been extensively employed for organics and nitrogen removal.324 Recent studies focused on understanding the fate and removal of antibiotics in SBR and showed effective removal of various antibiotics such as OFX: 65% (15 mg/L),238 SMX: 37% (67 μg/L),325 and 80% (10 mg/L),326 TMP: 21% (0.5 μg/L),327 and TET: 97% (250 μg/L)328 and 68% (10 mg/L).329 The antibiotics removal efficiency in both ASP and SBR was found to vary depending on the physical−chemical properties of antibiotics, the specific treatment process employed, 81 redox conditions, 327 SRT,330,331 hydraulic retention time (HRT),179,328 and temperature.30,95 HRT and SRT are the two critically important operating factors affecting antibiotics removal, and SRT > 15 days and HRT > 4−6 h are needed to achieve efficient removal of antibiotics.190,326,329 The aqueous phase removal efficiencies of TMP (initial concentration of 620 ng/ L) were 71%, 34%, and 19% at SRT of 20, 7, and 2 days, respectively, in laboratory-scale SBR studies.331 Longer SRT leads to enhanced microbial diversity and higher biomass concentration179 thereby facilitating the enrichment of slowgrowing microbes capable of degrading the antibiotics.331 The contact time required for the degradation of sulfonamide antibiotics was reported to be longer than 4−6 h, which is a typical HRT of conventional ASP.190 The removal of SMX and TMP in laboratory-scale SBR operating under different redox

ring, and the reaction is catalyzed by esterase encoded by the Ere gene family (Figure 3d).302 There are, however, a lack of studies on biodegradation pathways of MLs in anoxic and anaerobic (methanogenic and SRB) conditions. 4.2.5. β-Lactams. PEN G and AMX are the most widely used β-lactam antibiotics due to their high antimicrobial activity and low toxicity and cost.303 However, PEN G, AMX, and other β-lactam antibiotics are not widely detected in the environment, mainly due to their chemically unstable β-lactam ring, which is highly sensitive to environmental conditions (such as pH, temperature, etc.).304 For example, PEN G transformed to penillic acid (Figure S15A in Supporting Information) under acidic conditions, or penicilloic acid (Figure S15B in Supporting Information) and penilloic acid (Figure S15C in Supporting Information) under alkaline conditions, which were the main intermediates detected in the environment.305,306 PEN G could be transformed to penicilloic acid via hydrolysis of the β-lactam ring and subsequently degraded via decarboxylation (Figure S15 in Supporting Information).304 The hydrolysis of the β-lactam ring of PEN G and other β-lactam antibiotics can be catalyzed by βlactamases (such as serine-β-lactamases and metallo-βlactamases).302 Like PEN G, the degradation of AMX began with the hydrolysis of the β-lactam ring, which was then further converted to other intermediates. Amoxicillin diketopiperacine-2′,5′ and amoxilloic penicilloic acid diastereomers were the major intermediates detected in WWTPs (Figure 3e and Figure S16 in Supporting Information).307 It is apparent that the degradation of PEN G, AMX, and other β-lactam antibiotics in the environment is mainly initiated via the hydrolysis of the β-lactam ring.302 There are, however, limited studies on biodegradation mechanisms of β-lactam antibiotics in different biological sludge systems.

5. FATE AND REMOVAL OF ANTIBIOTICS IN ENGINEERED BIOLOGICAL TREATMENT SYSTEMS Most antibiotics could only be removed partially during the conventional wastewater/manure treatment/sludge digestion process, and they are subsequently being released into the environment through effluent discharge and manure/sludge disposal.308,309 Various studies investigated antibiotics removal efficacy of different biological wastewater treatment systems and reported a variable degree of removal with contradictory results, especially with respect to bioreactor operating conditions. In this section, we systematically synthesize information regarding the removal of different antibiotics in various biological treatment systems (conventional and advanced) with discussion on their removal efficiency, removal mechanisms, critical bioreactor operating conditions/parameters, and new advancements in bioreactor development for enhanced removal of antibiotics. The overview of different treatment systems for antibiotics removal is summarized in Table 2. The removal efficiency of antibiotics from the aqueous phase is calculated as the percentage reduction in concentration between influent and effluent.3,95,310,311 In engineered biological treatment systems, antibiotics can be removed from the aqueous phase via biodegradation or sorption onto the biological sludge.15,312 A few studies investigated these two mechanisms and provided the sorption and biodegradation contributions to the overall removal based on antibiotics concentration in aqueous and solid (sludge) phase.282,312−315 It is worth mentioning that the mass balance assessment of 7246


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biodegradation was found to be the major removal pathway for SAs and TMP.338,339 Removal of FQs, e.g., CIP and OFX, was mainly mediated by sorption onto soil matrix and wetland media, respectively.334,340 Antibiotics removal in CWs varies considerably depending upon the flow type and direction,333,338,339,341,342 wetland media type,333,334,339,340 plant species,336,339,340,343 hydraulic loading rate (HLR),339,340 temperature,95,339,341 and class of antibiotics and their concentration. The vertical subsurface flow (VSSF) CWs showed a higher antibiotic removal efficiency than HSSF and free water surface (FWS) flowCWs due to a better oxygenation and superior rhizosphere effect promoting rhizodegradation as well as biosorption.339,341,342 For instance, SMX removals of 76% and 13% were reported in VSSF-CWs and surface flow (SF)-CWs, respectively.342 Flow directions (up-flow and down-flow) influence the contact between antibiotics and wetland media, HRT, and redox conditions, which potentially affect the spatial distribution of microbial communities in CWs.333 Longer HRT maintained in vertical up-flow-CWs greatly contributed to higher antibiotics removal as compared to vertical down flowCWs.344 Higher removal of TMP in the up-flow VSSF-CW system was mainly attributed to anaerobic biodegradation as evident from low oxidation−reduction potential (ORP) (−192 to −258 mV).339 CW systems with wetland media (e.g., brickrubble, zeolite and vesuvianite) having a higher specific surface area, bed porosity, and smaller average pore size showed enhanced removal of antibiotics (such as FQs, TCs, SAs, and TMP) by promoting better microbial growth and providing longer HRT.333,334,339,340 The antibiotics (e.g., OFX, SMZ, and ROX) removal in the CW system planted with Umbrella palm (Cyperus alternifolius) was higher than those without the plant mainly due to plant uptake and the rhizosphere effect which enhanced microbial degradation.345 Seasonal variation also affected antibiotics removal, and higher removal efficiency of different antibiotics including SAs (SMZ, SMX and TMP), TCs (TET, CTC), MLs (ROX), and FQs (OFX and CIP) was reported in summer than in winter95,339,341 due to enhanced microbial activities at a higher temperature. Apart from poor microbial activity in winter, both desorption of substratebound antibiotic compounds and the potential cleavage of conjugates could contribute to the negative removal.95 Despite better performance of CWs, one of major limitations is that they typically operate at low HLR and thus require a large surface area.346 Hybrid systems consisting of CWs of different configurations connected in series could be operated at higher HLRs. For instance, a system consisting of vertical flow (VF) CW followed by horizontal flow (HF) CW and FWS flow-CW in series, exhibited good removal of antibiotics such as ENRX: 90% (initial concentration of 1.8 μg/L), DOXC: 58% (initial concentration of 3.7 μg/L), SMX: 30% (initial concentration of 2.5 μg/L), ERY: 10% (initial concentration of 1.9 μg/L) and lincomycin (LIN): 18% (initial concentration of 2.9 μg/L) in spite of operating at high HLR of 0.18 m/day (m3/m2/day). The higher removal is attributed to simultaneous occurrence of aerobic/anaerobic biodegradation, sorption, and photodegradation processes.346 However, further research is needed to gain insight into the removal pathways, effect of CW design, and operational parameters on the removal of diverse antibiotics in CW systems. 5.1.3. Anaerobic Reactor Systems. Anaerobic digestion (AD) is a matured technology and is most frequently used for

conditions: aerobic, sequential anoxic/aerobic, and microaerobic (DO concentration ∼0.3 mg/L) showed very low removal (85%) of antibiotics including SAs (sulfadimidine (SMD) and SMX), TCs (TET, CTC, OTC and DOXC), FQs (ENRX, CIP and NOFX), and MLs (tylosin (TYL)) at an average influent concentration of 44.8 μg/L in swine farm wastewater.332 SAs were mainly removed via biodegradation, while removals of TCs and FQs were contributed by both biosorption and biodegradation.332 5.1.2. Constructed Wetlands (CWs). CWs are engineered wastewater treatment systems and are effective in removing antibiotics from swine farm333−335 and municipal wastewaters.336−338 A few studies demonstrated comparable or higher removal efficiency of the selected antibiotics in CW systems over conventional ASP.95,336 Common reed (Phragmites australis)-based horizontal subsurface flow (HSSF) CW system exhibited higher aqueous phase removal efficiency than the conventional ASP (e.g., DOXC (71% in CW vs 61% in ASP at concentration of 0.18 μg/L), CLAR (31% in CW vs 18% in ASP at concentration of 0.25 μg/L), ERY (64% in CW vs 0% in ASP at concentration of 0.06 μg/L), SMX (87% CW vs 60% ASP at concentration of 0.26 μg/L), SDM (99% CW vs 53% ASP at concentration of 3.1 μg/L), and TMP (99% CW vs 14% ASP at a concentration of 0.1 μg/L)).336 Better removal could be attributed to the significantly longer HRT in the CW system (∼ 24 to 120 h) compared to conventional ASP (∼6−16 h)95 and combination of different processes, namely, biosorption onto soil or wetland media, plant uptake, and microbial degradation.333,334 Several studies examined the contribution of different removal pathways (biosorption on wetland media and soil, plant uptake, photolysis, and biodegradation) to the overall removal of antibiotics in CW systems,334,335,338−340 and 7247


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Environmental Science & Technology the treatment of animal manure and sewage sludge.347,348 Studies on antibiotics sorption and biodegradation during AD are very limited.349−352 AD was more effective than other sludge stabilization methods (e.g., aerobic digestion, anaerobic stabilization ponds) for the removal of antibiotics (e.g., CIP, NOFX, OFX, SMX, and TMP) from waste activated sludge.353 Significant reduction in antibiotic concentrations from swine manure and raw sewage sludge was obtained in laboratory- and pilot-scale digestion studies (Table 2).354,355 Biosorption on the digested sludge was the major removal pathway for FQs and TCs.350,351,354 Although a significant decrease in antibiotic concentration was observed both in laboratory- and pilot-scale AD systems, there are limited studies on degradation mechanisms and pathways of antibiotics along with distribution in aqueous phase and adsorbed antibiotics in full-scale manure digestion systems. Temperature is the most critical factor affecting the AD process, with higher antibiotics removal at mesosphilic (32−37 °C) and thermophilic (50−60 °C) conditions as compared to psychrophilic (20 days) maintained in granular reactor facilitated enhanced removal of antibiotics.407,409 The granular sludge SBR with SRT of 23−28 days demonstrated a higher removal of SMX (84%) compared to suspended sludge SBR (73%) with SRT of ∼10 days.409 The aqueous phase removal efficiencies of SMX and NOFX (initial concentration of 50−55 μg/L) were 78% and 84% (SRT = 30 days), 60% and 70% (SRT = 15 days), respectively.410 Few novel configurations including aerobic granular sludgebased MBRs (AGS-MBRs) were examined for their efficacy in antibiotics removal such as SMX and NOFX, with an average removal efficiency of 78% and 88%, respectively.410 More recently, application of granular sludge in anaerobic SBR was investigated for treatment of AZI containing wastewater. On the basis of mass balance analysis, biodegradation was found to be the major removal pathway for AZI with an average removal efficiency of 31%.411 Though granular systems exhibited high resistance to elevated antibiotics concentration without a significant impact on overall organic removal, few studies highlighted the influence of antibiotics on granular biomass properties and structural stability. Decrease in granular size and disintegration of granules was evident due to continuous exposure of antibiotics such as FQs (OFX, NOFX and CIP), TCs (TET), and BLAs (AMPI and CEFX) at higher concentrations.403,408,412,413 However, more studies are required to better understand the influence of different antibiotics on the stability and morphology of granules during long-term continuous operation of bioreactors. Additionally, studies on the use of granular sludge-based bioreactors in full-scale WWTPs for antibiotics removal are still lacking. 5.2.5. Sulfur-Mediated Biological Systems. Recently, sulfur-mediated biological systems, such as the novel sulfate reduction, autotrophic denitrification and nitrification integrated (SANI) process,414 flue gas desulfurization-SANI

(FGD-SANI),415 and denitrifying sulfur-assisted enhanced biological phosphorus removal (DS-EBPR) process416 have attracted significant attention due to their effectiveness in carbon, nitrogen, and phosphorus removals, and overall low biological sludge production, CO2 emission, and energy consumption.414,417 Moreover, an SRB-enriched system has been employed for the treatment of sulfate-laden pharmaceutical industry and saline municipal wastewater, and SRBs have shown high tolerance against pharmaceutical compounds.417−419 More recently, our studies showed that SMX and CIP could be effectively removed by SRB sludge in an anaerobic up-flow reactor system via adsorption and biotransformation with an aqueous phase removal of 34% (SMX concentration of 100 μg/L) and 60% (CIP concentration of 5000 μg/L).253,282 Biotransformation contributed to nearly 28% removal of CIP in an SRB sludge system.282 Importantly, SRBs (i.e., Desulfobacter) showed tolerance to a high concentration of antibiotics (CIP: 5000 μg/L), thereby opening up a potential of novel SRB-based biological process for antibiotics removal.203,253,282 However, further studies are needed to gain insights into antibiotic removal mechanisms, biodegradation pathways, dominant microbial species, and functional genes catalyzing their biodegradation in the SRBbased biological process.

6. FUTURE RESEARCH OUTLOOK The relative dominance of two removal pathways (i.e., adsorption and biodegradation) primarily depends on physical−chemical properties of the antibiotics and the operating conditions of WWTPs. There is a strong evidence that sorption and biodegradation of antibiotics differ significantly under different redox conditions (e.g., aerobic, anoxic, and anaerobic (methanogenic and sulfate-reducing)) employed in wastewater treatment processes. Variation in the redox environment can significantly influence the abundance and distribution of microbial communities, thereby resulting in different sorption and biodegradation mechanisms. Comparative studies especially correlating the biochemical characteristics of aerobic, anoxic, and anaerobic sludges to antibiotics adsorption are needed. Furthermore, studies on metabolic pathways of antibiotics removal and the microbial communities involved (especially, under anoxic and anaerobic redox conditions) are very limited. In-depth studies using molecular approaches such as metagenomics to reveal the abundance, diversity, and distribution of functional genes catalyzing the antibiotics biodegradation in mixed microbial system are needed. Little is known about the fate of largely unknown intermediates of antibiotics biodegradation under different redox conditions. The potential environmental toxicity of such intermediates should also be considered by coupling biological assays along with biodegradation studies. Studies on the role of EPS on adsorption of antibiotics and the associated complex interactions between EPS and the antibiotics are needed to better understand the fate of antibiotics in engineered biological wastewater treatment systems. Unlike EPS, SMPs are the organic byproducts of substrate metabolism with diverse functional groups, released into the aqueous phase, which may interact with antibiotics similar to EPS. However, SMP-bound antibiotics may still be present in the solution. Further studies are needed to gain insight into the role of SMPs on antibiotics removal, especially in membrane bioreactor systems. 7250


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Environmental Science & Technology Although advanced hybrid engineered biological wastewater treatment systems combining different redox conditions and biomass conformations are effective in removal of different antibiotic compounds, more in-depth investigations are required to evaluate their performance in full-scale WWTPs considering different operating parameters. Despite major advancements, biological wastewater treatment systems are only moderately effective in the removal of antibiotics. Thus, to further improve the antibiotic removal in WWTPs, there is need to upgrade the conventional biological treatment process by integrating with additional physical− chemical treatment technologies such as activated carbon adsorption, membrane (e.g., ultrafiltration, microfiltration, nanofiltration, reverse osmosis, and forward osmosis), and advanced oxidation processes (AOPs) (e.g., photo-Fenton, electro-Fenton, direct ozonation, sonochemical, and heterogeneous photocatalysis) among others.299,421,424−430 In recent years, forward osmosis has become more popular for removal of emerging trace organic contaminants as compared to other membrane-based separation technologies, as the process is less susceptible to fouling.425 With advances in material science, new adsorbents such as carbon nanotubes,431 graphenes, or graphene-based adsorbents (e.g., graphene oxide-coated biochar nanocomposites),432 and metal−organic frameworks (MOFs)433 could be more effective for enhanced removal of different antibiotics. More recently, nanotechnology has shown unprecedented opportunities in enhancing the removal of a wide variety of micropollutants in existing wastewater treatment processes.434,435 Significant progress has been made to enhance photocatalysis through nanotechnology-based approaches such as via surface modification and the use of nanocomposites.434 For example, the surface defects of titania could be eliminated by using porous adsorbents such as carbon nanotubes (CNTs) thereby resulting in a higher adsorption capacity of the catalyst due to an increased surface area and photocatalytic activity during the degradation of antibiotics.436 However, more comprehensive bench-scale followed by pilotscale studies are needed to thoroughly understand the fate and removal of antibiotics, and for successful integration of these advanced treatment processes into the existing WWTPs. Finally, future research needs to be directed toward developing unified models focusing on antibiotics transport and fate in full-scale WWTPs. The influence of different operational parameters on biotransformation, retransformation, and biosorption/desorption of antibiotics should be taken into consideration while developing the model. Thus, combining experimental works with modeling could offer a new opportunity for better understanding of the ultimate fate and transport of antibiotics in WWTPs.

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tion in aquaculture effluents, effluents from pharmaceutical industries, sewage sludge, surface and groundwater, adsorption coefficient (Kd) values, thermodynamic equations and parameters, EPS components and their concentration, binding constant (Kb) values; 16 figures highlighting different classes of antibiotics and their mode of action, antibiotics usage, occurrence (concentration profile) of antibiotics in domestic, hospital and livestock farm wastewater, antibiotics ionization forms, biotransformation pathways for different antibiotics (PDF)

AUTHOR INFORMATION

Corresponding Author

*E-mail: [email protected]. Phone: +(86) 20 39333161. Fax: +(86) 20 39333161. ORCID

Samir Kumar Khanal: 0000-0001-6680-5846 Hui Lu: 0000-0002-0084-9229 Author Contributions

A.S.O.: formulated the content and scope of the paper, involved in interpretation of all information, and prepared the first draft of the manuscript. Y.J.: involved in writing the biodegradation section and figures, and assisted in reviewing the manuscript. H.Z.: involved in preparation of the figures and assisted in writing the EPS section. S.K.K.: involved in planning the content of the paper and supervised the entire manuscript preparation. H.L.: developed the initial idea, involved in planning the content of the paper, and supervised manuscript preparation as a principal investigator. Author Contributions #

A.S.O. and Y.J. contributed equally.

Notes

The authors declare no competing financial interest.

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ACKNOWLEDGMENTS The study is being supported by the Natural Science Foundation of China (Nos. 51778643 and 51638005) and the Pearl River S&T Nova Program of Guangdong (201504281527416). The authors also acknowledge the support from Tip-top Scientific and Technical Innovative Youth Talents of Guangdong Special Support Program (No. 2016TQ03Z336).

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REFERENCES

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ASSOCIATED CONTENT

S Supporting Information *

The Supporting Information is available free of charge on the ACS Publications website at DOI: 10.1021/acs.est.9b01131. Details of physical−chemical properties of diverse classes of antibiotics, global consumption of antibiotics, information about potential sources contributing to antibiotic load into the environment, information about analysis and quantification of antibiotics and details of adsorption isotherms; 15 tables showing details of physical−chemical properties of selected target antibiotics, antibiotics consumption, antibiotics concentra7251


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