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Undoubtedly domestic Wastewater Treatment Plants (WWTPs) are not designed for the removal of some pollutants such as antibiotics. This chapter summarizes the occurrence and fate of six groups of the most widely used antibiotics (β-lactams, sulfonamides, quinolones, tetracyclines, macrolides, and others) in domestic WWTPs. The literature showed that the six groups of antibiotics have been frequently detected during wastewater treatment train (influent, primary treatment, secondary treatment, tertiary treatment, effluent, and sludge treatment) of domestic WWTPs. Also, it was clear that the main removal routes of antibiotics during sewage treatment of domestic WWTPs were adsorption, biodegradation, membrane filtration, and disinfection. Domestic WWTPs cannot remove most of the antibiotics which finally enter the environment through treated effluent and sludge.
Water Pollution Research Department, Environmental Research and Climate Changes Institute, National Research Centre, Dokki, Egypt
Majid M. Alruwaili
Projects and Technical Services, National Water Company Headquarter, Riyadh, Saudi Arabia
Fahad S. Aljohani
Projects and Technical Services, National Water Company Headquarter, Riyadh, Saudi Arabia
Shimaa R. Hamed
Microbial Biotechnology Department, Biotechnology Research Institute, National Research Centre, Dokki, Egypt
Samar Ragab
Biomedical Sciences Department, Center of Microbiology and Phage Therapy, Zewail City of Sciences and Technology, Giza, Egypt
*Address all correspondence to: alwasifyr2013@gmail.com
1. Introduction
Nowadays, environmental researchers have extended their focus beyond classic environmental contaminants such as pesticides, Polychlorinated Biphenyls (PCBs), and dioxins [1, 2]. Antibiotics are one of the new serious environmental contaminants. Antibiotics, a group of pharmaceuticals used as a medicine and growth promoter for both humans and animals, are considered a new serious environmental contaminant due to their continuous input into the environment and persistent presence [3, 4, 5].
Antibiotics are chemicals classified depending on their nature into three main categories: natural, semi-synthetic, and synthetic. Moreover, antibiotics can be classified depending on their mode of action into two categories: bactericidal (kill microorganisms) and bacteriostatic (impede microbial growth).
In developing countries, there is an increase in the consumption of antibiotics, without any prescription from physicians, to cure the different kinds of diseases that originate from improper general hygiene and poor sanitation systems. Also, antibiotics are used widely in animal farming for the protection of animal health to maintain the high demand for animal products [6].
Boyles et al. [7] reported that the high consumption of antibiotics by humans (households, hospitals, and industry) and in veterinary results in the increasing release of unchanged active ingredients and partially metabolized antibiotics into the sewer system (directly or indirectly) which consequently reach into the domestic wastewater treatment plants [8, 9, 10].
Conventional wastewater treatment plants cannot completely remove antibiotics and these antibiotics will finally contaminate the environment through effluent or sludge [11, 12]. Therefore, local wastewater treatment plants act as one of the main pathways for antibiotics to transfer into the environment [13].
Although antibiotics residues in water are very low (ng/L to μg/L), they still draw the researchers’ attention in the whole world since these antibiotics are the main source for the occurrence and transfer of Antibiotic-Resistant Genes (ARG) and Antibiotic-Resistant Bacteria (ARB) which have serious impacts on the environment [14, 15].
During last years, simultaneous detection of trace concentrations of antibiotics in wastewater and sludge samples is no longer difficult as a result of the invention of new detection methods such as liquid chromatography-mass spectrometry, solid-phase extraction, and ultra-performance liquid chromatography-mass spectrometry with the rapid development of analytical methods such as Solid Phase Extraction (SPE), High-Performance Liquid Chromatography Mass Spectrometry (HPLC-MS/MS) and Ultra-Performance Liquid Chromatography Mass Spectrometry (UPLC-MS/MS) [16, 17].
The occurrence of antibiotics in water environments such as groundwater and surface water was summarized by previous studies [18, 19, 20, 21, 22] as well as sediments, sludge, and soil [19]. Nevertheless, most of these studies focused only on the occurrence of antibiotics in different environments with little focus on the fate of antibiotics. For example, summarizing the removal of antibiotics in sediment, water, and soil environments, instead of the elimination of antibiotics in wastewater treatment plants [21]. Also, the elimination of Personal Care Products (PPCPs) and pharmaceuticals via biodegradation and other pathways in wastewater treatment plants, with limited content on antibiotics elimination [23].
Thus, in this review, we summarize the data and information on the occurrence and fate of antibiotics in wastewater treatment plants (WWTPs) to provide the overall profile of antibiotics concentrations in influent, treatment stages, sludge, and effluent of wastewater treatment plants, and to understand the elimination routes and fate of antibiotics in WWTPs.
2. Occurrence of antibiotics in the aquatic environment
Domestic wastewater treatment plants receive most of the used antibiotics through the sewer network, while the rest of the antibiotics are dumped directly into rivers and streams or escape as leachate from landfills. Figure 1 summarizes the introduction pathways of antibiotics into the aquatic environment.
Figure 1.
Pathways of antibiotics in the aquatic environment, WWTP: Wastewater treatment plant, DWTP: Drinking water treatment plant.
3. Occurrence of antibiotics in wastewater treatment plants (WWTPs)
Antibiotics can be classified using different ways such as their chemical structure. According to the chemical structure, there are 12 different classes of antibiotics (Table 1) such as β- lactams, aminoglycosides, macrolides, glycopeptides, oxazolidinones, sulfonamides, quinolones (fluoroquinolones), polymyxins, tetracyclines, streptogramins, and others such as chloramphenicol, thiamphenicol, lincomycin, trimethoprim, and clindamycin [5, 8]. Table 1 summarizes the chemical structure of the different 12 classes of antibiotics and their mode of action, mechanism of action as well as discovery dates.
Table 1.
Different classes of antibiotics and their mode of action and mechanism of action.
In this review, Tables 2–7 summarize data about the occurrence of the most common antibiotics (6 classes) in influent and effluent samples of wastewater treatment plants. These antibiotics include β-lactams, quinolones, sulfonamides, macrolides, tetracyclines, and others, while Table 8 summarizes the occurrence of these antibiotics in sludge generated from wastewater treatment plants.
Antibiotic
Concentration (ng/L)
Sampling type
Treatment technology
Influent
Effluent
Penicillin G
29.0
NA
Grab
Primary treatment
ND-10
ND-300
Composite
—
Penicillin V
20–13,800
ND–2000
Grab
Primary treatment
Amoxicillin
190–280
ND–30
24 h FPC
Membrane filtration/Reverse osmosis
1400–6940
ND-50
Composite
—
Ampicillin
NA
7
Grab
Activated sludge
Oxacillin
NA
<20
24 h FPC
Activated sludge + chlorination
Cloxacillin
ND–320
ND
Grab
Activated sludge
ND-4600
ND-700
Composite
—
Cefalexin
1200
980
Grab
Activated sludge + denitrification and nitrification
NA
1110–1410
24 h FPC
Chemically enhanced primary treatment
2800–64,000
ND-250
Composite
—
Cefotaxime
24
<12
Grab
Activated sludge + denitrification and nitrification
Figure 2 shows the typical structure and the treatment train of a domestic wastewater treatment plant. The treatment train begins consists of three stages including primary, secondary, and tertiary. The primary treatment stage contains preliminary treatment units (screens and sand & grit removal), an oil and grease removal unit, and primary settling tanks (not common in most WWTPs). The secondary treatment stage contains a biological treatment unit (aeration tank) using different methods such as activated sludge (AS) and trickling filters (TF), followed by secondary settling tanks (secondary clarifiers). The tertiary treatment stage contains a sand filtration unit followed by a disinfection unit using different technologies such as chlorination, ozonation, and ultraviolet (UV), and finally nutrient removal unit (for treated effluent which contains high concentrations of phosphorus and nitrogen).
Figure 2.
Typical structure of domestic wastewater treatment plant.
3.1 Presence of antibiotics in influent and effluent samples of WWTPs
3.1.1 β-Lactams
β-lactams are a group of antibiotics characterized by the presence of the β-lactam ring. The β-lactams ring is the main structure that gives the antibacterial activity for β-lactams. The variation in pharmacological properties depends on the differentiation in the side chains. β-lactams include two subclasses which are cephalosporins and penicillins [24].
The β-lactam antibiotics industry has annual sales of about $15 billion [25]; this makes up 65% of the total antibiotics market [26]. 6-Aminopenicillanic acid is the precursor to produce β-lactam antibiotics, which can be produced using free and immobilized penicillin G acylase [27]. Immobilized enzymes are preferred over the free enzyme for many reasons including product-free enzymes as reported by Elnashar [28].
In the samples collected from WWTPs (Table 2), some penicillins were detected such as ampicillin, penicillin G & V, amoxicillin, cloxacillin, and oxacillin. Penicillin V showed the highest concentrations (2000–2013,800 ng/L) in influent and effluent wastewater samples [29]. For cephalosporins, six types were detected including cefotaxime, cefalexin, cefaclor, cloxacillin, cefazolin, and cephradine [5, 30]. Generally, β-lactams, especially penicillins were rarely detected in domestic wastewater though they are used in great amounts [14, 31]. This may be attributed to the unstable nature of the β-lactam ring. The β-lactam ring can be broken by a popular bacterial enzyme group called β-lactamases [22] or can be broken through the chemical hydrolysis process [32].
3.1.2 Sulfonamides
Sulfonamides are used commonly since 1968 and consist of a large group of broad-spectrum antibiotics [5, 13, 33]. In WWTPs, 16 types of sulfonamides have been detected (Table 3). The most detected sulfonamides were sulfamethoxazole (5597–6000 ng/L) followed by sulfamethazine, sulphapyridine, and sulfadiazine, sequentially. N4-acetylsulfamethoxazole and many other N4-acetylated sulfonamides, dominant human metabolites of sulfonamides, were also detected in WWTPs and it was discovered that these metabolites can be transformed into their parent compounds [11, 34, 35].
3.1.3 Quinolones (fluoroquinolones)
Quinolones are a class of antibiotics characterized by the presence of quinolone branches as their basic structure. Fluoroquinolones are a subclass of quinolones that contains a fluorine-substituted central ring. Nalidixic acid was the first quinolone antibiotic discovered in the 1960s followed by newly developed four generations of quinolones [3, 35]. As a result of the universal extensive usage of quinolones, all generations of these antibiotics were detected worldwide in WWTPs as shown in Table 4. The most commonly detected quinolones in WWTPs were ciprofloxacin, norfloxacin, and ofloxacin [8, 12, 36]. The highest concentrations detected in influent and effluent samples were 4600 and 7870 ng/L for ciprofloxacin and ofloxacin, respectively [11, 37, 38].
3.1.4 Tetracyclines
Tetracyclines are composed of eight antibiotics (semisynthetic and natural) which inhibit the synthesis of bacterial proteins. Tetracyclines are widely used for human use, the poultry industry, and animal agriculture [39]. In WWTPs, five types of tetracyclines were detected in influent and effluent samples and tetracycline was the most detected one as shown in Table 5 [5, 14]. Doxycycline showed the highest concentration (2210 ng/L) in influent samples [40], while tetracycline showed the highest concentration (1420 ng/L) in effluent samples [37]. Tetracyclines were discovered in sludge samples with no or little biodegradation. Generally, as shown in Table 8, tetracyclines have a relatively rare presence in WWTPs because tetracyclines are used rarely by humans [41, 42].
3.1.5 Macrolides
Macrolides are a group of antibiotics characterized by the presence of a lactone ring that is substituted with alkyl, hydroxyl, and ketone groups, which inhibit the synthesis of bacterial proteins and are usually used as substitutes for penicillin [1, 34]. In WWTPs, six types of macrolides and one metabolite of erythromycin (erythromycin-H2O) were detected in influent and effluent samples. The most frequently detected macrolide was erythromycin-H2O followed by roxithromycin, clarithromycin, azithromycin, and tylosin in sequence as shown in Table 6. The lowest detected macrolides in sequence were erythromycin-H2O, roxithromycin, clarithromycin, azithromycin, and tylosin. Erythromycin-H2O showed the highest concentrations in the influent (10,025 mg/L) and in the effluent (4330 ng/L) wastewater samples [37].
3.1.6 Others
In this review, the other category of antibiotics consists of five different types including chloramphenicol, trimethoprim, lincomycin, clindamycin, and thiamphenicol. All these other antibiotics were detected in WWTPs, and trimethoprim was the most abundant and widely distributed one as shown in Table 7. Trimethoprim showed the highest concentrations in influent (7900 ng/L) and in effluent (3052 ng/L) samples, while Chloramphenicol and thiamphenicol showed the lowest concentrations (<4 to 1050 ng/L) in WWTPs [14, 41].
3.1.7 Summary
Despite that β-lactams are the most consumed antibiotic by humans, they were not detected frequently due to the unstable nature of β-lactams [21]. Many reasons affecting the significant variation of antibiotics’ concentrations in wastewater influents such as:
the consumption pattern of antibiotics, which may be different in different countries and even in the same country there is a great variation in consumption patterns of antibiotics [21].
different sampling procedures; grab/composite sampling results in a great variation in antibiotic concentrations since pharmaceutical loads reached WWTPs are different during the day [43].
hourly and seasonal fluctuation; seasonal variation of antibiotic consumption was high between summer and winter since antibiotic consumption is two times higher than consumption in summer which affects the concentrations of antibiotics in wastewater influent samples [44, 45]. Also, during winter, antibiotics in influent samples are diluted by rain [11, 46].
wastewater treatment plant scale, which means that higher concentrations of antibiotics in influent samples occurred when WWTP serves a low population and vice versa. Additionally, the concentrations of antibiotics varied in effluent depending on the different sampling methods and the applied wastewater treatment technology [5, 8].
3.2 Presence of antibiotics in sludge samples of WWTP
The analysis process for antibiotics and their transformations in sludge samples is a challenge since antibiotics in sludge have a low detectability rate as well as a low extraction pattern [47]. Therefore, the studies on the detection and presence of antibiotics in sludge are much less than those on wastewater [48]. The results showed the presence of five different classes of antibiotics in both activated and digested sludge samples as shown in Table 8.
The most abundant antibiotics were ciprofloxacin (4.8 mg/kg) in digested sludge [40] and norfloxacin (2.7 mg/kg) in activated sludge [14, 49]. Quinolones were detected mostly at the level mg/kg, while most tetracyclines, macrolides, sulfonamides, and others were detected in lower levels (μg/kg) [11]. Despite the wide usage of β-lactams antibiotics in veterinary and human medicine, all sludge samples all over the world showed no presence of β-lactams antibiotics which is mainly attributed to their poor adsorption onto sludge and their unstable characteristics [37].
Antibiotics in WWTPs can be removed through major pathways including biodegradation, adsorption, membrane separation, and disinfection. Furthermore, there are other pathways for the removal of antibiotics including photolysis, volatilization, and hydrolysis which were eliminated since they have an inconsiderable role in the reduction of antibiotics from WWTPs. For example, β- lactams antibiotics are not stable because of the presence of β- lactam ring which easily can be hydrolyzed. Thus, β-lactams will be hydrolyzed before reaching the WWTPs.
Additionally, some researchers reported that β- lactams have a relatively long half-life due to hydrolysis at neutral pH values (same as in WWTPs) such as more than 5 days for amoxicillin [50] and 52 h for meropenem [51]. Moreover, although β-lactams can be hydrolyzed in WWTP, the contribution of the hydrolysis process for the removal of antibiotics is useless because the wastewater treatment process has a relatively short hydraulic retention time (8–20 h).
In addition to that, there are some antibiotics such as amoxicillin which were degraded by sunlight-photolysis or UV-photolysis [50], macrolides [52], quinolones [53], and tetracyclines [54, 55]. However, this degradation process has a minor significance because wastewater has high concentrations of suspended solids which inhibit the deep penetration of sunlight or UV [5, 8]. Besides, the effect of the photolysis process can be neglected since the hydraulic retention time of WWTPs is much lower than the half-life of most antibiotics in wastewater.
4.2 Antibiotics transformation and fate in conventional WWTP
4.2.1 Primary treatment units
The primary treatment stage in WWTPs consists mainly of screens and primary settling tanks. In some WWTPs, some coagulant chemicals are added in primary treatment units such as ferric ion salts, aluminum salts, or polymers called CEPT (Chemically Enhanced Primary Treatment) [56]. Many previous studies reported that the primary treatment stage in WWTPs has no significant removal for different types of antibiotics including clarithromycin, sulfamethoxazole, ofloxacin, cefalexin, erythromycin, amoxicillin, clindamycin, and cefaclor [13].
Nevertheless, other studies reported that the chemically enhanced primary treatment (CEPT) process can significantly improve the removal efficiencies of some antibiotics such as norfloxacin (67.7%), ofloxacin (55.2%), erythromycin (44.8%), sulfamethoxazole (64.0%), and roxithromycin (76.3%). This effect is due to the destroying effect of coagulants on the chemical chains of some antibiotics [11].
4.2.2 Biological treatment units
Clearly, in biological treatment units, biodegradation and adsorption processes are the main pathways for the transformation of antibiotics in WWTPs. According to the classification of antibiotics, their transformation and fate in biological treatment units can be summarized as the following:
4.2.2.1 β-Lactams
Despite that β-lactams are the most consumed antibiotics for humans and animals, they have not been detected frequently in WWTPs [14, 57], thus there was not much-published data about the fate and transformation of β-lactams in WWTPs. Junker et al. [57] studied the fate of some 14C-labeled antibiotics (benzylpenicillin and ceftriaxone) in the activated sludge process. The results showed that ceftriaxone was not totally mineralized, whereas only about 25% of benzylpenicillin was mineralized. The same results were obtained using biodegradability tests (closed bottle tests method; CBT) at much higher concentrations of β-lactams antibiotics since ceftriaxone was kept unchanged whereas benzylpenicillin was biodegraded up to 27% [58].
The differences in biodegradability between β-lactams antibiotics may be due to the differences in their chemical structures because of diverse side chains [24, 59]. Andreozzi et al. [50] performed a standard batch experiment to study the fate of amoxicillin in the activated sludge process. The results proved that the adsorption and biodegradation processes were responsible for the removal and transformation of amoxicillin.
4.2.2.2 Sulfonamides
Most researchers who studied sulfonamides in WWTPs focused on sulfamethoxazole and N4-acetylsulfamethoxazole (a metabolite of sulfamethoxazole). Batt et al. [41] and Pérez et al. [60] found that sulfonamides were biodegraded to a certain degree (low removal efficiency) in the wastewater biological treatment stage. Sulfonamides were removed during the biological treatment process with an average removal efficiency of 25% [13, 38], as well as sulfamethoxazole showed poor removal efficiency of 20% [42, 61].
Also, some studies reported the resistance of sulfonamides to different treatment processes during wastewater treatment [11, 12, 62]. Nevertheless, some other researchers reported the relatively high removal efficiency of sulfamethoxazole like 55% [34], 56% [63], 66% [64], 67% [65], and 74% [63].
The significant variation in removal efficiencies of sulfonamides during the biological treatment process can be attributed to the following reasons: first, the transformation of some metabolites such as N4-acetylsulfamethoxazole to the parent molecule (sulfamethoxazole) in the influent. Second, the mentioned removal efficiencies depended on grab or composite samples (24 h), which cannot reflect the whole treatment process [12, 66].
Thus, to avoid the previous limitations, some researchers used well-controlled laboratory reactors for studying the fate of sulfonamides and their removal pathway during the activated sludge treatment process [8, 67]. The biodegradation process of three sulfonamides at low concentrations (20 μg/L) using activated sludge reactors was studied by Pérez et al. [60] and they reported that the biodegradation process was so efficient and was able to remove the three sulfonamides for 3 days. Less than 26% of the initial antibiotics’ concentrations were present by the third day, whereas by the tenth day, the removal efficiency increased up to 93%. In addition, it was reported that some microorganisms can utilize sulfamethoxazole as a carbon and/or nitrogen source after 3 days lag phase [68]. Despite these studies proving that sulfonamides can be biodegraded, the biodegradation process takes a long time than the usual hydraulic retention time of the biological treatment process at WWTPs.
4.2.2.3 Quinolones
Batt et al. [41] and Xu et al. [12] reported that adsorption followed by biodegradation are the main pathways for the removal of quinolones during biological treatment stages at WWTPs. The removal efficiencies for norfloxacin, ofloxacin, and ciprofloxacin were 87–100%, 75–77%, and 85%, respectively [69, 70]. The adsorption mechanism of quinolones by sludge depends on electrostatic interactions between particles rather than hydrophobic forces [21, 42].
4.2.2.4 Tetracyclines
In the biological treatment process at WWTPs, adsorption is considered the main mechanism for the removal of tetracyclines [41, 71, 72]. Tetracycline (10 μg/L) was removed (up to >95%) rapidly through an adsorption mechanism during 6 h inside activated sludge units. Also, two lab-scale Sequencing Batch Reactors (SBR) were utilized to stimulate the activated sludge process (biological treatment), in these SBRs, the effect of SRT (Sludge Retention Time) and HRT (Hydraulic Retention Time) on transformation and fate of tetracycline were studied (66). The results showed that the removal efficiency of tetracycline in phase 1 (SRT = 10 days; HRT = 24 h) was 86.4 ± 8.7% and phase 2 (SRT = 10 days; HRT = 7.4 h) was 85.1 ± 5.4%, while in phase 3 (SRT = 3 days; HRT = 7.4 h) was 78.4 ± 7.1%. In phase 3, it was clear that the removal efficiency of tetracycline decreased by a reduction in SRT, which indicated that more tetracycline can be adsorbed by old sludge. In addition to that, it was reported that ferrous chloride could enhance the removal of tetracycline through precipitation due to the strong complexation between tetracyclines and ferrous ions [41].
4.2.2.5 Macrolides
All previous studies have indicated that all macrolides were not significantly eliminated, even at low concentrations, during the biological treatment process at WWTPs [10, 67].
4.2.2.6 Trimethoprim
Many studies indicated that trimethoprim was not adsorbed during the Activated Sludge (AS) process [41, 60]. The studies also proved that trimethoprim cannot easily biodegrade during AS with a short sludge retention time [11, 57]. However, Pérez et al. [60] reported that trimethoprim was completely degraded by nitrifying activated sludge with long SRT within 3 days. The nitrifying bacteria present in the nitrifying activated sludge are responsible for trimethoprim degradation since it was noticed that when the activity of nitrifying bacteria is inhibited, the elimination efficiency of trimethoprim decreased from 70 to 25% [42, 73].
4.2.3 Digestion tank
A little number of studies were carried out to study the transformation and fate of antibiotics in digestion tanks because the digestion process is not applied at most of the WWTPs all over the world. Zhang and Bing [5] conducted a two-stage anaerobic sludge digester (SRT = 30 days) to study the fate of fluoroquinolones (norfloxacin and ciprofloxacin) and they reported the stability of these fluoroquinolones. The same results were reported by Lindberg et al. [66] that ciprofloxacin and norfloxacin showed no significant removal under mesophilic sludge digesters (38°C). In contrast, Du et al. [14] utilized anaerobic mesophilic sludge digesters to study the stability of trimethoprim and sulfamethoxazole and they reported the instability of trimethoprim and sulfamethoxazole.
This may be due to that both trimethoprim and sulfamethoxazole have no significant amounts in digested sludge while their concentrations are high in activated sludge. Gartiser et al. [74] studied the biodegradation of nine antibiotics under an anaerobic digestion process (35 ± 1°C) and they reported that the biodegradation process was inefficient for all nine antibiotics except in the case of benzylpenicillin which was biodegraded after 40 days lag phase.
4.3 Antibiotics transformation and fate in advanced treatment processes
4.3.1 Filtration
To improve the quality of treated effluent, some wastewater treatment plants apply advanced treatment units such as Membrane Filtration (MF) or Sand Filtration (SF). During sand filtration, the removal efficiencies of trimethoprim and clarithromycin were 60 and 15%, respectively [8, 34]. Nakada et al. [75] also reported the same results for trimethoprim since the removal efficiency was 55.2% in the sand filtration process. However, sulfamethoxazole and sulphapyridine showed lower removal efficiencies of 26.9 and 14.6%, respectively after the sand filtration process.
Moreover, clarithromycin, azithromycin, roxithromycin, and erythromycin-H2O showed no removal at all during the sand filtration process which may be due to the presence of highly diverse and effective biofilm on the SF particles [75]. Watkinson et al. [38] reported that about 43 and 94% of total antibiotics were removed by the microfiltration process and reverse osmosis, respectively via eliminating the particles that adsorbed antibiotics on them. Also, the Nanofiltration (NF) process increased the removal efficiencies of antibiotics up to more than 95% for antibiotics such as tetracyclines [42, 76].
4.3.2 Disinfection
Table 9 summarizes the transformation and fate of eight classes of antibiotics after disinfection units at WWTPs. The disinfectants reacted fast with the antibiotics and the removal occurred after 1–27 min. In most studies, ozonation was applied for disinfection purposes. The ozonation oxidized antibiotics either directly by ozone (O3) or by hydroxyl radicals (•OH). Hydroxyl radicals are generated due to the decay of ozone. Each ozone and hydroxyl radicals have different oxidation mechanisms since O3 is selective and usually attacks the special functional groups (such as the aromatic structure or a C〓C double bond), whereas hydroxyl radicals are non-selective and react with many types of moieties [75].
Antibiotic
Disinfectant
pH
Performance
Contact time (min)
Removal efficiency (%)
β-lactams
Penicillin G
Ozone
7.7
1
–70
Cefalexin
Ozone
7.7
1
>99
Sulfonamides
Sulfamethoxazole
Free available chlorine
7.3
1
−95
Sulphapyridine
Ozone
7.95
27
93.9
Fluoroquinolones
Ciprofloxacin
Free available chlorine
7.7
1
>99
Enrofloxacin
Ozone
7.7
1
>99
Macrolides
Azithromycin
Ozone
7.95
27
92.6
Erythromycin-H2O
Ozone
7.95
27
88.7
Clarithromycin
Ozone
7.2
18
>76
Tylosin
Ozone
7.7
1
>99
Roxithromycin
Ozone
7–7.5
8-40-4
>90–99
Tetracyclines
Tetracycline
Ozone
7.7
1
>99
Amikacin
Ozone
7.7
1
−25
Vancomycin
Ozone
7.7
1
>99
Others
Trimethoprim
Free available chlorine.
8.1
10
−100
Ozone
7.7
1
>99
Lincomycin
Ozone
5.5
2
100
Table 9.
Transformation and fate results of antibiotics in the disinfection process.
In addition, Cha et al. [77] studied the ozonation of 14 antibiotics and found that only 4 antibiotics including cephalexin, penicillin G, N(4)-acetylsulfamethoxazole, and amikacin were oxidized by hydroxyl radicals and the other 10 antibiotics reacted mainly with O3. Moreover, by using other disinfectants such as hypochlorous acid (HOCl), some antibiotics such as trimethoprim showed no degradation, while sulfamethoxazole was degraded after a reaction with HOCl [78]. Liu et al. [10] reported that there was a significant variation in reaction rates of combined chlorine and free available chlorine with antibiotics in wastewater.
This review provides insight into the occurrence and fate of antibiotics in domestic wastewater treatment plants. Data was collected about the occurrence of the most widely six groups of antibiotics used for human cure including β-lactams, sulfonamides, quinolones, tetracyclines, macrolides, and others in wastewater and sludge samples of wastewater treatment plants. All previously mentioned groups of antibiotics were detected in wastewater and sludge samples with varied concentrations during the different treatment stages. It was clear that most of the wastewater treatment plants do not have the ability to fully remove these antibiotics. The main removal mechanisms of these antibiotics were biodegradation, adsorption, membrane filtration, and disinfection.
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Written By
Raed S. Al-Wasify, Majid M. Alruwaili, Fahad S. Aljohani, Shimaa R. Hamed and Samar Ragab
Submitted: 24 May 2023Reviewed: 29 May 2023Published: 23 August 2023
Open access peer-reviewed chapter
