Current Natural Degradation and Artificial Intervention Removal Techniques for Antibiotics in the Aquatic Environment: A Review

Abstract

The extensive use of antibiotics as essential medications in contemporary healthcare has resulted in significant amounts of these drugs entering the environment, both in original and metabolic forms, which presents serious ecological and health hazards. This paper examines the natural processes that break down antibiotics in water, including photolysis, hydrolysis, and biodegradation. It also discusses advancements in artificial degradation technologies, such as advanced oxidation processes (AOPs), physicochemical methods, ionizing radiation degradation, artificial wetland technology, microalgae technology, microbial electrochemical systems, and innovative catalysts. While current technologies demonstrate promising potential for use, they encounter challenges related to the catalyst stability, cost, and ecological safety. Future research should focus on optimizing degradation methods and creating efficient, sustainable multi-technology systems, such as the photocatalysis–membrane filtration coupling system; the ultrasound–Fenton–artificial wetland synergistic system; the electrochemical–biodegradation combined system; and the microbial fuel cell (MFC)–photocatalysis synergistic system, to tackle the complexities of antibiotic pollution in the environment.

1. Introduction

Antibiotics are a crucial class of medications in modern healthcare, commonly used to prevent and treat infectious diseases in both humans and animals by either destroying bacteria or inhibiting their growth and reproduction [1]. Since Alexander Fleming first discovered the natural antibiotic penicillin in 1929, hundreds of antibiotics have been developed for clinical use through the extraction of natural compounds and the molecular modification of existing antibiotics. Currently, there are ten major classes of antibiotics categorized by their chemical structures: penicillins, macrolides, cephalosporins, quinolones, beta-lactams, lincosamides, tetracyclines, sulfonamides, glycopeptides, and aminoglycosides. Among the antibiotic classes commonly detected in water are tetracyclines, beta-lactams, fluoroquinolones, sulfonamides, and macrolides [2]. Different types of antibiotics exert specific effects on various species of bacteria, with the primary mechanisms being (1) the targeted inhibition of protein synthesis through interactions with ribosomal subunits (e.g., tetracycline, chloramphenicol, and aminoglycosides), (2) the disruption of cell wall synthesis (e.g., β-lactams), (3) interference with nucleic acid processes (e.g., rifampicin and fluoroquinolones), (4) the disruption of metabolic pathways (e.g., folate analogs such as sulfonamides), and (5) the alteration of the bacterial membrane structure (e.g., polymyxins). The main classifications and mechanisms of the action of antibiotics are listed in

Table 1. Main classification and resistance mechanisms of antibiotics.

Antibiotic Class	Antibiotic Examples	Primary Target
Protein Synthesis Inhibitors
Aminoglycosides	Tobramycin, Gentamicin, Kanamycin	30S ribosomal subunit (causes tRNA misreading)
Tetracyclines	Oxytetracycline, Doxycycline, Tetracycline	30S ribosomal subunit (blocks aminoacyl-tRNA binding)
Macrolides	Erythromycin, Spiramycin, Azithromycin	50S ribosomal subunit (inhibits elongation step)
Phenicols	Chloramphenicol, Thiamphenicol, Florfenicol	50S ribosomal subunit (inhibits elongation step)
Cell Wall Synthesis Inhibitors
β-Lactams	Penicillin, Cephalosporins	Penicillin-binding proteins (disrupts cell wall synthesis)
Glycopeptides	Ampicillin, Vancomycin	Peptidoglycan subunits (inhibits cell wall biosynthesis)
DNA Synthesis Inhibitors
Quinolones	Ciprofloxacin, Levofloxacin	Topoisomerase II (DNA gyrase), Topoisomerase IV
Sulfonamides	Sulfamethoxazole, Sulfadiazine	Competitive inhibitor of DHPS (dihydropteroate synthase) in folate synthesis
RNA Synthesis Inhibitors
Rifamycins	Rifampin, Rifaximin	DNA-dependent RNA polymerase
Cell Membrane Disruptors
Polymyxins	Polymyxin B, Polymyxin E (Colistin)	Bacterial lipopolysaccharide layer (increases membrane permeability)

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Antibiotics are extensively utilized in the management of infectious diseases in both humans and animals, as well as being employed as feed additives to promote growth in livestock and poultry. The research conducted by Judit Lienert indicates that only a small percentage of ingested antibiotics are metabolized and absorbed by the body during their use. The majority of these substances (50–90%) and their primary metabolites are not fully metabolized by humans or animals, leading to their excretion through urine and feces. This excreted material retains significant bacteriostatic activity and ultimately enters the natural environment [3]. The excretion rates are 98% for sulfamethoxazole, 82% for amoxicillin, 60% for ciprofloxacin, and less than 50% for norfloxacin and erythromycin. More concerningly, practices such as overuse, inadequate regulation, and improper handling have led to elevated levels of antibiotic residues being detected in various aquatic environments, with some even contaminating drinking water [4]. The main sources of antibiotic contamination in the environment are as follows: (1) Households—the vast majority of antibiotics used by humans (50–90%) are excreted in urine or feces as parent compounds or metabolite intermediates, which end up in municipal wastewater treatment plants (wastewater treatment plants (WWTPs)). (2) Hospitals—the major sources of large quantities of antibiotics in municipal wastewater, dominated by drugs originating from patients that are not metabolized and absorbed and excreted through urine (55–80%) and feces (4–30%). (3) Livestock and poultry farming—antibiotics are also widely used in animal husbandry (cattle, pigs, and poultry) and aquaculture; due to the excessive use of antibiotics in the process of livestock and poultry production and poor metabolic absorption in animals, a large number of antibiotics and their metabolite residues are transferred into the feces. (4) Aquaculture—the large-scale use of antibiotics in centralized aquaculture has led to the accumulation of large quantities of antibiotic residues in surface water and sediments near the farms. (5) Antibiotic production plants—antibiotic production plants generate large quantities of pharmaceutical wastewater containing high concentrations of antibiotics during their operation, and their antibiotic load in the environment is significantly higher than that of traditional sources, such as municipal and hospital wastewater (Figure 1).

According to reports, extremely high concentrations of antibiotics are frequently detected in various sewage treatment plants and areas where manure is discharged. Liu Yi summarized that the highest median concentration of sulfamethoxazole in the Liao River in China reached 60.9 ng/L; the highest median concentration of ciprofloxacin in the Yellow River was 65.9 ng/L. Additionally, the highest median concentration of erythromycin in the sediments of the Pearl River was 72.6 ng/g [5]. Booth et al. reported that the average concentrations of ciprofloxacin, sulfanilamide, tetracycline, and trimethoprim in industrial wastewater reached values as high as 3548.6 μg/L, 18,416.8 μg/L, 23,119.0 μg/L, and 3078.7 μg/L, respectively. In addition to the wastewater sources of these antibiotics, residues of various antibiotics have also been detected in surface water and groundwater worldwide, particularly at elevated levels in surface water near livestock farms [6]. For example, Le et al. found that the concentration of sulfamethoxazole in surface water near aquaculture shrimp ponds reached as high as 5.6 mg/L [7]. In the surface water around livestock farms in Jiangsu Province, China, sulfonamide antibiotics were detected at levels of 560–4660 ng/L, and tetracycline antibiotics were found at levels of 810–2420 ng/L [8]. Jiang et al. also noted that suburban areas along the river have intensive livestock activities and that the overall antibiotic contamination is more severe in the suburbs than in the cities [9]. Changes in antibiotic concentrations in the aquatic environment are consistent with changes in farm patterns, with residual concentrations of antibiotics in farms varying seasonally. Typically, detectable frequencies and average concentrations were higher in winter than in summer [10].

Once introduced into aquatic ecosystems, antibiotics can exert toxic effects on the structure, growth, respiration, and physiological activity of aquatic microorganisms, including bacteria, cyanobacteria, algae, and daphnia [11,12]. Algae serve as the foundation of the food chain, and even a minor decline in algal populations can destabilize existing food chains and food webs, disrupting ecosystems at all levels [13]. Tetracycline and sulfonamide antibiotics have been demonstrated to inhibit algal growth by impacting chloroplast replication, transcription, translation, and metabolic pathways in algae [14]. More concerningly, antibiotics can bioaccumulate in fish muscle after prolonged exposure. Consuming aquatic organisms that contain these accumulated antibiotics may increase health risks for humans [15].

In addition, the presence of antibiotics in aquatic environments exerts selective pressure on native bacterial communities. This not only poses a threat to aquatic organisms but also contributes to the development of antibiotic-resistant bacteria (ARB) and resistance genes (ARGs) [16]. The study notes that ARGs are transmitted between different bacteria through horizontal gene transfer (HGT), which accelerates the proliferation of cross-species resistant organisms. This phenomenon could facilitate the transfer of ARGs from aquatic environments to human and animal pathogens, significantly reducing the effectiveness of antibiotics for treating infectious diseases and posing serious risks to both human and animal health [17]. The Centers for Disease Control and Prevention reports that approximately two million people are infected with antibiotic-resistant bacteria each year, resulting in at least 23,000 deaths. If antibiotic-resistant infections are not effectively addressed, it is projected that 10 million people could die globally each year by 2050 [18]. The management and treatment of antibiotic residues in water bodies are crucial for global health. This paper reviews the mechanisms of antibiotic degradation in natural environments and specific techniques for antibiotic removal, including hybrid methods, to assess the advancements in current technologies and ultimately enhance the effectiveness of antibiotic removal.

2. The Degradation of Antibiotics in the Natural Environment

2.1. Photolysis and Hydrolysis

The natural degradation of antibiotics in the environment primarily occurs through photolysis and hydrolysis. The photolysis process depends on the absorption of specific wavelengths of light by antibiotic molecules, which initiates photochemical reactions, including both direct and indirect photolysis. The process of direct photolysis refers to the direct absorption of sunlight by antibiotic molecules, particularly the photon energy in the ultraviolet–visible light spectrum, which triggers the breaking of chemical bonds or structural rearrangements, leading to degradation. The direct photolysis pathway is evident in the photochemical degradation of drug molecules that contain photon-absorbing groups, such as quinolones [19]. Indirect photolysis requires an excitation donor, known as a photosensitizer, to transfer excitation energy to an acceptor molecule, such as an antibiotic, in order to facilitate a photochemical transformation. This can include reactions such as reactive oxygen species (ROS, such as ¹O₂, O₂⁻, •OH, H₂O₂, etc.)-induced photo-oxidative degradation and hydrated electron-induced photo-oxidative degradation. However, the efficiency of these processes is significantly influenced by a variety of environmental conditions. Yen-Ching Lin discovered that quinolone antibiotics, such as ciprofloxacin, exhibit a higher affinity for soils and sediments. Additionally, oxidizers and catalysts, including MnO₂, Fe (III), and Al (III), which are commonly found in the natural environment, may significantly influence their oxidation and indirect photolysis. Among them, Fe³⁺ and Al³⁺ form complexes with the carboxyl or carbonyl groups of ciprofloxacin, while MnO₂, a metal oxide, fixes ciprofloxacin molecules through surface adsorption, thereby affecting its photostability. Additionally, Fe³⁺ and MnO₂, as strong oxidants, can directly oxidize organic matter or generate reactive oxygen species through electron transfer when exposed to promote indirect photolysis. Conversely, Al³⁺ alters the molecular conformation of ciprofloxacin through complexation or electrostatic adsorption, thereby inhibiting its photolysis. Other factors, such as humic substances, clay types, and pH, can also impact photolysis in the solid phase. At the same time, TiO₂ can generate strong oxidizing electron–hole pairs (e⁻-h⁺) under ultraviolet (UV) light, producing hydroxyl radicals (•OH) and superoxide radicals (O₂⁻). It has high photocatalytic activity, is low-cost, highly stable, and non-toxic, making it the most commonly used catalyst in photolysis [20]. Other factors, such as the type of clay, pH levels, and the presence of humic substances, can also impact photolysis in solid phases [21]. For example, dissolved humic acid (HA) contains various functional groups, including carboxyl groups, phenolic hydroxyl groups, carbonyl groups, amino groups, and methoxy groups. These negatively charged functional groups can influence van der Waals forces and hydrogen bonding during the adsorption of organic pollutants in the soil [22]. Furthermore, the effect of dissolved organic matter (DOM) on the environmental behavior of antibiotics varies significantly. For instance, DOM derived from biochar enhances the adsorption of sulfamethoxazole (SMX) while inhibiting the adsorption of chloramphenicol onto biochar [23]. The pH level directly affects the migration and transformation of sulfonamides (SAs) in the environment by altering their chemical state. At lower pH levels, sulfadiazine (SDZ) predominantly exists in a cationic form, leading many positively charged SAs to electrostatically bind to adsorption sites on the soil surface, which significantly increases the adsorption rate of SDZ. As the pH rises, the electrostatic repulsion between SDZ and soil materials diminishes the adsorption capacity [24].

In addition, the presence of dissolved organic matter (DOM), such as humic acid (HA), significantly influences the photodegradation process of antibiotics. DOM is the primary form and active component of natural organic matter in aquatic environments. Research has demonstrated that DOM plays a dual role in the photolysis of antibiotics. It can either undergo direct photolysis or function as a photosensitizer to facilitate indirect photolysis. The latter involves the generation of reactive oxygen species (ROS), with photochemically reactive DOM being the main source of ROS in surface waters. Antibiotic molecules can experience rapid oxidative degradation in the presence of ROS [25]. DOM can undergo direct photolysis or function as a photosensitizer to facilitate indirect photolysis. This indirect photolysis involves the generation of reactive oxygen species (ROS), with photochemically reactive DOM serving as the primary source of ROS in surface water. Antibiotic molecules can experience rapid oxidative degradation when exposed to ROS [25]. However, there may also be situations that inhibit light degradation [26,27,28]. This phenomenon can be attributed to the light-protective capabilities of dissolved organic matter (DOM), which create a shielding effect, quench photoproducts of reactive intermediates (PPRIs), and regulate the inherent excited state characteristics of antibiotic molecules. As a result, this process inhibits the photodegradation of antibiotics [29]. Humic acid (HA) has a dual role in the photodegradation of tetracycline (TC). At low concentrations of HA (<6 mg/L), it acts as a photosensitizer; under simulated sunlight/xenon lamp irradiation, HA absorbs light energy to generate excited triplet state ³HA* and produces reactive oxygen species (ROS) such as ¹O₂, O₂⁻, and •OH, which indirectly oxidize TC and promote its photodegradation. At high concentrations of HA (>6 mg/L), it competes for photon absorption, reducing the direct photodegradation efficiency of TC and quenching the ROS generated by TC itself (e.g., O₂⁻), thereby inhibiting TC photodegradation [30]. However, sometimes the promoting effect of HA on TC photodegradation is not obvious, possibly because the ROS generated by HA are quickly quenched by TC or other components, leading to a significant inhibition of TC photodegradation by HA (5 mg/L) at a pH of 8.0–9.0 [31].

The hydrolysis pathway of antibiotics is effective for certain types of pollutants, particularly in the degradation of esters and amides. The hydrolysis of β-lactam compounds typically occurs through the opening of the β-lactam ring and the cleavage of the amide bond. For instance, the hydrolysis of cephalexin primarily results from the rupture of the β-lactam bond and the positional alterations of the branched amide bond, which render the hydrolyzed form of cephalexin more amenable to biodegradation by microalgae [32]. In addition, pH levels and temperature can significantly influence the hydrolysis of antibiotics. For instance, research conducted by Shannon M. Mitchell assessed the hydrolysis conditions of β-lactams and discovered that the hydrolysis rates of ampicillin, cefazolin, and cefoxitin were positively correlated with temperature. Furthermore, the hydrolysis rate in alkaline environments was markedly higher than in acidic and neutral pH conditions [33].

2.2. Biodegradation

Microbial degradation is one of the important biological pathways for the removal of antibiotics from the environment. The degradation process is relatively complex and can degrade antibiotics through direct or indirect mechanisms, offering advantages such as high efficiency, environmental friendliness, and low costs.

The direct actions of the microbial degradation of antibiotics primarily include co-metabolism and enzyme-catalyzed reactions, such as hydrolysis and redox reactions. Co-metabolism, in particular, involves the breakdown of antibiotics into substances that can be utilized by microorganisms through non-specific degrading enzymes. This process requires the addition of carbon and nitrogen sources essential for microbial growth [34,35]. In enzyme-catalyzed reactions, the degrading enzymes produced by antibiotic-resistant bacteria primarily include the following: β-lactamase [36], aminoglycoside-modifying enzymes [37], macrolide inactivating enzymes [38], and chloramphenicol acetyltransferase [39]. They degrade antibiotics by modifying or hydrolyzing their molecular structures, resulting in the breakdown of these compounds. Research has found that both fungi and bacteria may be involved in the direct degradation of antibiotics. Bacillus cereus P41 is capable of completely degrading the β-lactam antibiotic cefotaxime in a tryptone soybean broth medium at 35 °C under anaerobic conditions [40]. The genus G1 of the yellow rod bacteria can degrade 80% of the macrolide antibiotic ampicillin in an MH culture medium at a pH of 7, using avermectin as the sole carbon and nitrogen source [41]. Under the conditions of a liquid culture medium containing 50 mg·L⁻¹ of oxytetracycline (OTC) and 3% malt extract, the rough-skinned side ear completely degraded the tetracycline antibiotic oxytetracycline [42].

The indirect mechanism of the microbial degradation of antibiotics involves microorganisms promoting the removal of antibiotics by altering environmental conditions or producing non-specific substances. This process includes microbial adsorption and redox reactions. Bioadsorption refers to the phenomenon in which substances are adsorbed onto the surfaces of organisms through covalent, electrostatic, or intermolecular interactions, such as utilizing activated sludge as a carrier for antibiotics. Research has demonstrated that activated sludge can provide adsorption sites for zwitterionic ions, with tetracycline (TC) ions binding to the activated sludge through cation exchange and electrostatic attraction mechanisms. The competition for adsorption sites between low-valent cations and TC ions can diminish the adsorption of TC by activated sludge [43]. During the process of antibiotic removal, microorganisms produce extracellular polymeric substances (EPSs), which are key components responsible for antibiotic adsorption. These substances act as a reservoir, allowing antibiotics to embed within the EPS matrix and accumulate, thereby enabling microorganisms to access growth materials more effectively. For instance, the EPS of microalgae has been demonstrated for the first time to play a significant role in bioadsorption, with the maximum absorption of ciprofloxacin (CIP) by the EPS being comparable to the total bioadsorption [44]. Redox reactions refer to the metabolic processes of microorganisms that produce electron donors or free radicals, which trigger the chemical degradation of antibiotics. Dissimilatory iron reduction (DIR) can induce the formation of iron redox species, thereby promoting metabolic activities and stimulating a series of beneficial reactions, such as biological Fenton reactions. Shewanella is one of the earliest isolated and thoroughly studied dissimilatory iron-reducing bacteria (DIRB), recognized for its ability to drive the Fenton reaction under neutral conditions that are conducive to its respiratory metabolism. The DIRB-driven Bio-Fenton process utilizes DIRB to reduce Fe³⁺ to Fe²⁺ through electron transfer, thereby enhancing the traditional Fenton reaction. DIRB also produces H₂O₂ as a by-product of its metabolic activity, which participates in the generation of •OH, thus degrading organic pollutants, including antibiotics [45].

However, most isolated strains have a single function and cannot adapt to real environments. In actual antibiotic degradation, high-diversity microbial communities can synergistically degrade antibiotics through various metabolic pathways (such as oxidation, reduction, and hydrolysis), achieving a higher degradation effect than single strains. Currently, the use of microbial communities to degrade fluoroquinolone antibiotics has been practically applied; for example, the microbial community in water filters can degrade ciprofloxacin in the water environment, achieving a removal rate of 89% after 28 days. In the degrading bacterial community, strains of the genera Rhizobium and Bacillus dominate [46]. The mixed microbial community AMQD4, composed of Providencia vermicola, Brevundimonas diminuta, etc., can degrade 56.8% of gentamicin (GEN), which is significantly higher than the degradation efficiency of single strains [47]. The community containing sulfate-reducing bacteria can degrade 84.2% of sulfamethoxazole, while the degradation efficiency of single strains is usually lower [48].

In recent years, research has revealed the potential risks of antibiotic resistance genes (ARGs) and antibiotic-resistant bacteria (ARB) carrying ARGs in the microbial degradation process. As a new class of pollutants, ARGs can replicate within microorganisms and be passed on to offspring through vertical transfer; they can also spread between different species mediated by mobile genetic elements (MGEs), facilitating gene transfer among bacterial communities [49]. What is even more concerning is that ARGs can migrate between different media and even transfer into the food web, posing a threat to ecological safety and human health [50]. Research has shown that antibiotic-resistant bacteria in drinking water can enter the human body through oral ingestion. Although diseases may not occur in the short term, infections can happen when the bacteria accumulate to a certain level and the body’s immune system is weakened [51]. In 2010, the superbug containing the NDM-1 gene was first discovered in the human body in New Delhi, India. This superbug can develop resistance to multiple antibiotics simultaneously, leading to human deaths due to a lack of effective treatments [52]. The serotype Escherichia coli 0104:H4 carrying the β-lactamase gene caused over 4000 infections of bloody diarrhea in Germany, leading to at least 850 cases of hemolytic uremic syndrome and 50 deaths [53]. Li, B. revealed the widespread distribution of antibiotic resistance genes (ARGs) in the environment and their co-occurrence with degrading strains through a metagenomic analysis, indicating that degrading strains may serve as potential vectors for the spread of ARGs [54]. Therefore, future research should further optimize microbial degradation strategies or utilize microbial community regulation techniques to enhance the degradation efficiency, while also assessing their ecological safety.

3. Artificial Intervention in the Degradation of Antibiotics

3.1. Advanced Oxidation Processes (AOPs)

Advanced oxidation processes (AOPs) are a new type of technology that degrade organic pollutants by generating strong oxidizing radicals, such as hydroxyl radicals (•OH), sulfate radicals (SO₄⁻•), singlet oxygen (O₂), and superoxide radicals (O₂⁻). Among these reactive oxidants, hydroxyl radical-based and sulfate radical-based advanced oxidation technologies are currently the most widely researched and applied, and they are the two most important categories [55,56]. Transition metals (both homogeneous and heterogeneous), ultraviolet light, ultrasound, conductive electrons, carbon catalysts, etc., activate oxidants (such as hydrogen peroxide (H₂O₂), peroxymonosulfate (PMS), peroxydisulfate (PDS), etc.) to produce strong oxidizing radicals for the degradation of organic pollutants. Typical advanced oxidation technologies based on •OH, such as Fenton, Fenton-like, catalytic ozone, photocatalysis processes and combinations or derived processes of these technologies, generally have advantages such as a simple operation, high degradation efficiency, and low cost and are widely used in the research of antibiotic degradation [57].

In the Fenton process, Fe²⁺/Fe³⁺ activates H₂O₂ to produce •OH, which oxidizes and degrades pollutants. •OH has a very high redox potential (2.8V) and can oxidize most pollutants into small molecular compounds, mineralizing them into CO₂ through in situ-generated ROS. Due to its simplicity and efficiency, it is widely used in the degradation of persistent pollutants [58]. For example, Affam et al. optimized the Fenton reaction to degrade the persistent antibiotic pollutants amoxicillin and cloxacillin using a response surface methodology [59]. Elmolla et al. achieved the oxidative degradation of amoxicillin, ampicillin, and cloxacillin in aqueous solutions using the Fenton method [60], and these results indicate that this method is effective for treating these three antibiotics [61].

According to the different forms of catalysts present and the reaction mechanisms, the Fenton reaction can be divided into homogeneous Fenton and heterogeneous Fenton. The advantages and disadvantages of homogeneous and heterogeneous catalysts are shown in the table below (

Table 2. The advantages and disadvantages of homogeneous and heterogeneous catalysts.

Comparison Item	Homogeneous Catalyst	Heterogeneous Catalyst
Advantages	1. High activity: direct contact with reactants enables fast reaction rates (Fe2+ completely degrades amoxicillin in 12 min at pH 3 [62]). 2. Mild reaction conditions: efficient reactions at room temperature and pressure [63].	1. Broad pH tolerance: some catalysts (e.g., Fe3O4/montmorillonite) function at neutral pH [64]. 2. Recyclability: magnetic materials (e.g., Fe3O4) can be separated via magnetic fields [65]. 3. High stability: e.g., FeS2 retains activity after repeated cycles [66].
Disadvantages	1. pH limitation: requires acidic conditions (pH 2–3), forms iron sludge in neutral environment [67]. 2. Difficult recovery: catalyst separation is challenging, leading to secondary pollution [68]. 3. Toxic by-products: may generate harmful intermediates [69].	1. Complex synthesis: requires carrier modification (e.g., graphene), increasing costs [70]. 2. Metal leaching risk: long-term use may release Fe ions [64].

).

The rate of the Fenton reaction is significantly affected by the pH, Fe²⁺ concentration, and H₂O₂ dosage, and sometimes the reaction may produce intermediates with a higher toxicity than the parent compounds, which are the major drawbacks of the Fenton reaction. Research indicates that the efficiency of the Fenton reaction is highest under acidic conditions (pH 2–4). However, under neutral or alkaline conditions, Fe²⁺ tends to precipitate, resulting in a loss of catalytic activity and the generation of a substantial amount of iron sludge, which contributes to secondary pollution [71]. This necessitates additional treatments, thereby increasing costs and environmental burdens. For example, He et al. noted that traditional homogeneous Fenton reactions encounter a significant precipitation of iron ions at pH levels greater than four. However, by developing a magnetic Fe₃S₄ catalyst, the effective pH range can be extended to neutral (pH = 7), although the catalytic efficiency remains lower than that observed under acidic conditions [72]. At the same time, to address the issue of iron sludge, Du et al. developed hydrogel microspheres loaded with iron ions, which reduce the generation of iron sludge through solid-phase catalysis. However, after five cycles of use, the catalytic efficiency decreased by 20%, indicating limitations in effectively resolving this issue [73]. Additionally, the utilization rate of hydrogen peroxide (H₂O₂) is low, and excessive additions can lead to increased costs and potentially trigger side reactions, such as hydroxyl radicals (•OH) being quenched by H₂O₂ [74]. This presents another issue that must be addressed in the Fenton reaction. Tang et al. improved the utilization rate to 70% by the in situ generation of H₂O₂ through a dual-cathode electro-Fenton system; however, the further optimization of the current efficiency is still necessary to reduce the energy consumption [75]. Bugueño-Carrasco et al. combined solar energy with the electro-Fenton system to reduce the amount of H₂O₂ added, but the efficiency significantly decreased on rainy days [76]. Additionally, if the intermediate products of the reaction are toxic, further treatment is required. For example, Dai et al. found that the short-chain carboxylic acids produced during the degradation of sulfamethoxazole have a persistent toxicity to Escherichia coli, necessitating an extension of the reaction time to achieve complete mineralization [77]; Zhu et al. pointed out that the degradation products of ofloxacin actually increased the toxicity to green algae, requiring the integration of a biological treatment to reduce risks [78].

To broaden the application range of Fenton technology, Fenton-like technologies have gradually become a popular research direction. Fenton-like reactions utilize other transition metals or solid-phase catalysts to replace divalent iron ions in catalyzing H₂O₂, generating free radicals that oxidatively degrade pollutants. Among these, transition metals (such as Fe, Co, Cu, and Ni) are of particular interest due to their low cost and favorable structural properties that facilitate regulation and modification [79]. Fixing metal ions in solid matrix materials to form heterogeneous reaction systems can improve the difficulty of the catalyst recovery and reduce sludge production, while also broadening the pH range. For example, Cu-based catalysts are considered one of the most important non-iron catalysts. The reaction rate of Cu⁺ with H₂O₂ (1.0 × 10⁻⁴ M⁻¹·s⁻¹) is faster than that of Fe²⁺ (76 M⁻¹·s⁻¹) [80]. Additionally, the reduction of Cu²⁺ is easier than that of Fe³⁺ because the standard redox potential of Cu²⁺/Cu⁺ is lower. At a neutral pH, Cu²⁺ can exist in the form of [Cu(H₂O)₆]²⁺, which allows Cu-based catalysts to exhibit a better pH tolerance and higher catalytic activity [81]. Furthermore, in a study by Li-Yuan Dai, solid-phase catalyst Fe₃O₄ nanoparticles were encapsulated in the lumen of halloysite nanotubes (HNTs) through electrostatic interactions, forming a core–shell structured nano-reactor system that achieved a degradation rate of nearly 100% for 50 mg/L of AF within 120 min, significantly outperforming traditional heterogeneous Fenton catalysts (such as pure Fe₃O₄ NPs or Fe₃O₄/HNTs composites) [82].

In recent years, research has focused on improving the efficiency of Fenton and Fenton-like reactions through catalyst modification and process optimization. Currently, common improved Fenton technologies include electro-Fenton technology, photo-Fenton technology, microwave Fenton technology, and ultrasound Fenton technology. For example, under acidic conditions in the electro-Fenton process, the anode generates a small amount of hydroxyl radicals (•OH) through the oxidation of water, while the cathode generates hydrogen peroxide (H₂O₂) in situ by reducing oxygen (O₂). At this time, under the influence of the electric field, a potential difference is formed on the surface of conductive or adsorptive particle materials, creating microelectrodes that increase the electrochemical reaction surface area and shorten the migration distance of reactants, thereby enhancing the degradation effect of target pollutants. In the photocatalytic Fenton reaction, titanium dioxide (TiO₂) is typically used as a catalyst, absorbing visible light and generating electron–hole pairs. When light strikes the surface of the photocatalyst, the photogenerated electrons and holes can participate in redox reactions. The holes have strong oxidative capabilities and can oxidize iron ions (Fe²⁺) to generate high-valent iron ions (Fe³⁺), while the photogenerated electrons can reduce hydrogen peroxide (H₂O₂) to generate hydroxyl radicals (•OH) [83]. In addition, in microwave Fenton technology, microwave radiation rapidly and uniformly transfers energy to the catalysts and organic pollutants in the reaction system, raising the temperature of the reaction system [84]. This increases the reaction rate and the generation of free radicals, promoting the degradation of organic pollutants. Ultrasonic Fenton technology, on the other hand, utilizes powerful acoustic vibrations generated by ultrasound in liquids, causing the precipitation and re-dissolution of dissolved gases in the liquid, thereby increasing the surface area of the gas–liquid interface [85]. This allows more oxygen to come into contact with the catalysts and organic pollutants during the Fenton reaction, increasing the generation of free radicals and the likelihood of reactions. Additionally, when the tiny bubbles generated by cavitation collapse, the high-temperature and high-pressure localized areas formed can further promote the generation of free radicals [86], accelerating the degradation process of organic pollutants in the Fenton catalytic oxidation reaction.

In addition, researchers are continuously developing emerging advanced oxidation processes (AOPs). For example, Priyadarshini reviewed AOPs based on peroxymonosulfate (PMS) and persulfate (PS), catalytic ozonation processes, cavitation-based AOPs, gamma radiation AOPs, and electrochemical AOPs. These processes generate highly reactive free radicals, such as hydroxyl radicals (OH) and sulfate radicals (SO₄²⁻), which can partially or completely mineralize emerging pollutants in wastewater. This research provides valuable insights for the practical application of AOPs in wastewater treatment, highlights the potential and challenges of these emerging technologies, and offers specific recommendations for future research directions [87].

3.2. Supercritical Water Oxidation

Supercritical water oxidation (SCWO) technology is an advanced oxidation method designed for the efficient treatment of challenging organic pollutants, which is particularly effective in degrading emerging contaminants such as antibiotics. Supercritical water, characterized by temperatures exceeding 374.15 °C and pressures above 22.1 MPa, exhibits unique physicochemical properties, including a low dielectric constant, high diffusivity, and strong solubility. These properties facilitate the rapid mineralization of organic substances into carbon dioxide (CO₂) and H₂O [88]. In recent years, the accumulation of antibiotics in the environment and their potential threats to ecology and public health have made supercritical water oxidation (SCWO) technology a focal point of research due to its efficiency and environmental sustainability.

The primary mechanisms by which SCWO degrades antibiotics include free radical oxidation, ring-opening reactions, and the cleavage of the central carbon chain. For instance, through a gas chromatography–mass spectrometry (GC/MS) analysis Gong et al. discovered that, in the absence of oxidants, the degradation products of tetracycline were predominantly tricyclic and bicyclic compounds. However, the addition of hydrogen peroxide (H₂O₂) led to the formation of smaller molecular chain compounds, such as acetamide, indicating that the oxidant enhanced mineralization [89]. Similarly, Stavbar et al. noted that the degradation of amoxicillin involves the cleavage of the β-lactam ring and the attack of hydroxyl radicals, ultimately resulting in the production of short-chain carboxylic acids and ammonium (NH₄) [90].

The degradation efficiency of antibiotics in supercritical water oxidation (SCWO) is significantly influenced by factors such as the temperature, oxidant concentration, reaction time, and initial pollutant concentration. Research indicates that elevated temperatures and the addition of oxidants can substantially enhance the degradation rate. For instance, Stavbar et al. treated wastewater containing amoxicillin and ciprofloxacin in a continuous flow reactor and discovered that at 500 °C with an oxidant coefficient of n = 1, the degradation rates of both antibiotics exceeded 98% [90]. Similarly, Gong et al. degraded tetracycline hydrochloride (TC) in a batch quartz tube reactor and found that at 400 °C, the addition of H₂O₂ (n = 1) resulted in the complete degradation of TC within 100 s [89]. Furthermore, the effect of the initial concentration on degradation efficiency is characterized by the principle that “low concentration leads to a high removal rate, while high concentration results in a high removal amount.” Dias et al. treated amoxicillin in pharmaceutical wastewater and achieved a total organic carbon (TOC) removal rate of 78% under optimized conditions (682 °C, n = 1), with hydrogen gas being the primary product in the gas phase (37.68%). This indicates that SCWO has the potential for both pollutant degradation and energy recovery [91].

The kinetics of the supercritical water oxidation (SCWO) degradation of antibiotics generally follows a first-order reaction model, and the addition of oxidants can alter the activation energy. For instance, a study by Gong et al. on the kinetics of tetracycline (TC) demonstrates that the activation energy without oxidants is 6.14 kJ/mol, which increases to 25.66 kJ/mol upon the addition of hydrogen peroxide (H₂O₂). This indicates that the oxidant facilitates a more comprehensive degradation pathway [89]. Additionally, Chen et al. elucidated the degradation mechanism of amoxicillin in supercritical water through molecular dynamics simulations, suggesting that radical attack and ring-opening reactions are the primary pathway [92]. Furthermore, Top et al. achieved a removal rate from 72% to 99.9% for nine types of pharmaceuticals when treating hospital wastewater at 550 °C (n = 2) conditions, with paroxetine exhibiting the fastest degradation, which is consistent with first-order kinetics [93].

3.3. Physical Chemistry Methods

3.3.1. Adsorption Method

Adsorption is a widely utilized method for the removal of antibiotics from water. Based on different adsorption mechanisms, it can be categorized into physical adsorption and chemical adsorption. This method involves the development of functional materials with a porous structure and effective adsorption performance, such as activated sludge, nanoadsorbents, surface-modified zeolites, biochar, etc. These materials purify water by adsorbing emerging pollutants, including antibiotics, through mechanisms such as electrostatic attraction, van der Waals forces, intermolecular forces, hydrogen bonding, and ion exchange. This technology is widely adopted and effectively utilized in the field of water treatment, both domestically and internationally, due to its advantages, including the simple equipment, ease of operation, mild reaction conditions, and cost-effectiveness, as well as its environmental friendliness.

The adsorption of antibiotics in activated sludge primarily involves hydrophobic and electrostatic interactions. Hydrophobic interactions occur when non-polar groups in antibiotic molecules bind to lipids or hydrophobic organic matter, such as extracellular polymeric substances (EPSs) present in sludge, through van der Waals forces. The logarithm of the partition coefficient (log P_ow) between octanol, which simulates biological membranes and organic matter, and water serves as a reference indicator. A log P_ow greater than 3.2 indicates a high hydrophobicity, facilitating the adsorption onto organic matter in sludge, including lipids and EPSs. Conversely, a log P_ow less than 0 indicates hydrophilicity, as exemplified by tetracycline, which has a log P_ow of −1.3. Although tetracycline exhibits low hydrophobicity, its adsorption can be enhanced by forming complexes with divalent metal ions, such as Ca²⁺ and Mg²⁺, or with EPSs in the sludge [94]. In activated sludge systems, the adsorption behavior of fluoroquinolone antibiotics, such as ciprofloxacin and norfloxacin, is highly dependent on the pH, which is closely related to the pKa values of the ionizable groups in their molecular structures. For instance, fluoroquinolones exhibit amphoteric properties, containing ionizable groups such as a carboxyl (-COOH, pKa ≈ 6) and amino (-NH2, pKa ≈ 8). When the system pH is below the pKa of the carboxyl group (approximately 6.0), the amino group in the molecule becomes protonated (-NH3+), resulting in the entire molecule carrying a positive charge. Since the surfaces of microorganisms and organic matter in activated sludge are typically negatively charged, a strong electrostatic attraction occurs between the two [95].

In addition, the high specific surface area of the nanoadsorbent demonstrates an excellent adsorption capacity for antibiotics. Yoon et al. reviewed the comprehensive research advancements in recent years regarding graphene-based nanoadsorbents for the removal of various antibiotics [96]. They discussed the application value of these nanoadsorbents in water and wastewater treatments, particularly summarizing their adsorption capabilities for different antibiotics. The authors evaluated the effects of key water quality parameters, such as the pH, temperature, and ionic strength, on the adsorption efficiency of various antibiotics. They conducted a comprehensive analysis of the factors influencing the adsorption mechanisms of nanoadsorbents and discussed their potential for regeneration and reusability.

Biochar, as a porous carbon material obtained from the pyrolysis of biomass, has a rich variety of surface functional groups and shows significant advantages in the adsorption of pollutants, including antibiotics, heavy metals, and organic pollutants. Dai et al. summarized the sources and production methods of biochar, highlighted the current research status regarding the removal of antibiotics, explained the mechanisms of antibiotic adsorption, introduced relevant adsorption parameters, and outlined regeneration methods and applications in engineering research. Finally, they analyzed the benefits of these applications and described the prospects for future development [97]. In recent years, research has indicated that Sargassum can be used to produce a new type of zeolite-like algal biochar (KSBC). The KSBC, doped with N, O, S, Al, and Si, exhibits zeolite-like properties, including good porosity, a high specific surface area (1137.60 m²/g), and a large number of oxygen-containing functional groups. Adsorption experiments were designed, and the results showed that the maximum adsorption capacities of ciprofloxacin and tetracycline on KSBC were 352.936 mg/g and 265.385 mg/g [98], respectively, providing new insights for the development of biochar.

3.3.2. Ultrafiltration Membrane Technology

Generally speaking, membranes can be classified into two categories: low-pressure membranes, which include microfiltration and ultrafiltration membranes, and high-pressure membranes, which encompass nanofiltration and reverse osmosis membranes. This classification is based on the molecular weight of the retained substances and the pore size of the membranes. The pore size of ultrafiltration membranes ranges from 0.05 μm to 1 nanometer, placing them between microfiltration membranes and nanofiltration membranes, and they are typically asymmetric membranes [99,100]. They effectively retain colloids and large molecular substances, such as viruses, bacteria, and natural organic matter, based on the size of the contaminants, allowing small molecular substances to pass through the membrane layer [101].

In membrane separation technology, ultrafiltration membrane technology has been widely applied in various large-scale industrial applications and water treatment processes [102] due to (1) its ability to produce a high-quality effluent; (2) its cost-effectiveness in terms of capital costs and low-pressure operating costs, which makes it suitable for large-scale applications; and (3) its high potential for macromolecule removal, a lower footprint, and low energy consumption [103]. For example, in the treatment of pharmaceutical wastewater, ultrafiltration can be used as a pretreatment unit to remove macromolecular organic substances, such as proteins and polysaccharides, thereby reducing the pollution load on subsequent nanofiltration or reverse osmosis processes [104].

Despite the numerous advantages of ultrafiltration membrane technology, its application in large-scale water treatment still encounters challenges, such as membrane fouling and difficulties in managing concentrates. During operation, ultrafiltration membranes are susceptible to contamination from organic matter, colloids, and microorganisms, which adhere to the membrane surface during the filtration process, forming a cake layer or embedding within the membrane pores. This contamination leads to a reduction in the membrane’s separation efficiency and results in membrane fouling. Therefore, to achieve a stable operation of the membrane separation technology and increase the efficiency of membrane materials, Hairong Yu [105] prepared ultrafiltration membranes by physically blending graphene oxide (GO) with polyvinylidene fluoride (PVDF). Due to the numerous hydrophilic groups on the surface of GO, a hydrated layer forms on the hydrophilic surface, hindering the adsorption of hydrophobic contaminants, such as proteins, on the membrane surface. Additionally, GO, as a hydrophilic additive, accelerates the exchange rate between the coagulation bath (water/ethanol) and the solvent (DMAc), resulting in faster phase separation, the formation of a wider finger-like pore structure, and improvements in the surface charge effects and mechanical strength. These multiple factors work together to enhance the membrane’s anti-fouling performance. In addition, Safarpour [106] prepared a novel ultrafiltration membrane made of a polyvinylidene fluoride (PVDF) mixed matrix containing a reduced graphene oxide/titanium dioxide (rGO/TiO₂) nanocomposite using a phase transformation method. Due to the high hydrophilicity of the rGO/TiO₂ nanocomposite, the rGO/TiO₂/PVDF membrane is more hydrophilic than the bare PVDF membrane, and the pure water flux increased by 54.9%. The membrane’s anti-fouling performance and flux recovery rate were significantly improved, with the best anti-fouling capability observed at a loading of 0.05 wt% rGO/TiO₂. In addition, the ultrafiltration process produces a concentrated solution with a high concentration of antibiotics, which may cause secondary pollution if discharged directly. Nasrollahi et al. pointed out that such concentrated solutions require further treatment, such as advanced oxidation or activated carbon adsorption, which increases the complexity and cost of the process [104].

Therefore, ultrafiltration membrane technology holds significant potential for large-scale water treatment, particularly as a pretreatment step or in combination with other technologies. However, challenges such as membrane fouling and concentrate treatments must be addressed. Future advancements will necessitate innovations in materials and process optimization to further improve the performance and cost-effectiveness of ultrafiltration membranes.

3.4. Ionizing Radiation Degradation

Ionizing radiation refers to radiation—such as electromagnetic waves or particle streams—that possesses sufficient energy to remove electrons from atoms or molecules in a substance, thereby generating ions. Examples of ionizing radiation include gamma rays, X-rays, and electron beams. This type of radiation is increasingly recognized as a promising alternative for degrading various toxic organic pollutants, including chlorophenols [107], nitrophenols [108], and antibiotics [109], which has attracted significant attention in many countries. The advantages of ionizing radiation technology include the absence of the need to add any chemical substances, an insensitivity to color and suspended particles, and a good permeability in aqueous matrices. Additionally, this technology can degrade difficult-to-degrade compounds in situ through reactive species generated during the water radiolysis process [110]. Radiolytic degradation primarily involves the production of strong oxidizing agents (•OH) and strong reducing agents (e-aq, H•) in water under ionizing radiation, which can effectively eliminate a wide range of pollutants. Currently, research is being conducted on the use of gamma rays or electron beam irradiation as advanced oxidation processes to remove antibiotics. For instance, Alsager et al. studied the degradation effects of γ-rays on amoxicillin, doxycycline, and ciprofloxacin. They found that at an initial concentration of 50 μM, a radiation dose of 7 kGy could achieve a removal rate exceeding 90%, with the degradation process adhering to a pseudo-first-order kinetic model [111]. Similarly, Wang et al. employed 30 kGy electron beams to degrade antibiotic residues in medical infusion bottles, achieving degradation rates of 97.02%, 97.61%, and 96.87% for amoxicillin, ofloxacin, and cephalexin, respectively [83]. Additionally, Liu et al. degraded ciprofloxacin (initial concentration of 20 mg/L) through γ-irradiation, achieving a removal rate of 88.35% at a dose of 400 Gy, and noted that lower concentration antibiotics were more readily degraded [112]. These studies demonstrate that ionizing radiation possesses a significant degradation capability for various types of antibiotics, with the degradation efficiency closely linked to the irradiation dose and initial concentration.

The efficiency of ionizing radiation in degrading antibiotics is influenced by various factors, including pH, inorganic anions, and organic matter. The pH level affects the degradation process by altering the generation ratio of free radicals. For instance, Guo et al. found that the degradation rate of sulfanilamide under acidic conditions at a pH of 3 (93.11%) was higher than that under alkaline conditions at a pH of 11 (87.30%), which is attributed to the more favorable generation of hydroxyl radicals (•OH) in acidic environments [113]. Conversely, research by Wang et al. indicated that alkaline conditions at a pH of 9 are more conducive to the electron beam degradation of antibiotics such as amoxicillin, suggesting that the optimal pH range for different antibiotics may vary based on their molecular structures [83]. Additionally, inorganic anions like carbonate (CO₃²⁻) and nitrate (NO₃⁻), as well as organic matter such as humic acid, can compete with target antibiotics for free radicals, thereby inhibiting the degradation efficiency. For example, Alsager et al. noted that the presence of CO₃²⁻ and NO₃⁻ significantly reduced the degradation rate of antibiotics [111], while humic acid decreased the removal rate of ciprofloxacin from 99.6% to 47.2% by consuming •OH [112].

Ionizing radiation technology has gradually transitioned from laboratory research to practical applications. For instance, in 2017, Jinhua, China, established its first factory for treating industrial wastewater using electron beam technology, showcasing the industrial potential of this approach [114]. To enhance the degradation efficiency in complex matrices, researchers often integrate ionizing radiation with other processes. For example, Rivas-Ortiz and colleagues combined γ-irradiation with Fe²⁺ /H₂O₂, resulting in an increase in the mineralization rate of sulfanilamide from 27.4% to 73.3% [115]. Furthermore, ionizing radiation can function as either a pretreatment or post-treatment step in biological treatment processes to improve the biodegradability of wastewater. Wang and Wang discovered that combining γ-irradiation with a biological treatment could elevate the total organic carbon removal rate of carbamazepine from 26.5% to 79.3%, underscoring the benefits of integrated processes [116].

3.5. New Type of Catalyst

Unlike the various techniques for degrading the antibiotics mentioned above, this section summarizes several novel catalysts developed in recent years that can significantly enhance both the degradation rate and the degree of mineralization of the antibiotics. These catalysts can be integrated with various degradation technologies and applied in the antibiotic degradation process. The methods involve enhancing the activity of redox reactions, expanding the range of light absorption, or providing additional active sites. The following are several key application scenarios of these new catalysts and their mechanisms of action, along with a summary (

Table 3. A summary of the new catalysts.

Catalyst Type	Representative Materials	Degradation Mechanism	Advantages	Limitations
Nanophotocatalysts	TiO2, ZnO, g-C3N4	Efficient separation of photogenerated electron–hole pairs and increased surface active sites. Heterojunctions (e.g., ZnO/TiO2) enhance charge separation; doping (N, C) broadens light absorption range and improves charge separation efficiency.	High catalytic activity; low cost.	Prone to e−/h+ recombination; some materials only respond to UV light; difficult to recover nanoparticles.
MOFs	MIL101(Fe), ZIF-8, UiO-66	Adsorption (electrostatic, π-π, hydrophobic, acid–base interactions); photocatalytic generation of radicals (e.g., O2− and •OH) for pollutant degradation.	Ultra-high specific surface area (strong adsorption capacity); adjustable pore structure; multifunctional active sites.	Poor water stability (some MOFs are prone to hydrolysis); high synthesis cost; Low separation efficiency of photogenerated carriers.
Carbon-based Single-Atom Catalyst	Ba/g-C3N4, Ag/mpg-C3N4	Single-atom metals (e.g., Ba and Co) optimize carrier band structure (e.g., g-C3N4) to enhance light absorption.	Atomic-level utilization (100%); significantly improved reaction kinetics.	Complex preparation process (requires precise control of single-atom dispersion); high cost.

).

3.5.1. Nano Photocatalyst

Catalysts such as TiO₂, ZnO, and g-C₃N₄ significantly enhance the efficiency of photocatalytic degradation by facilitating the effective separation of photogenerated electron–hole pairs and increasing the number of active surface sites. For instance, Min Li [117] prepared ZnO/TiO₂ heterojunction nanofilms using a magnetron sputtering method, employing a stepwise annealing strategy. The ZnO layer was annealed for 3 h, while the TiO₂ layer was annealed for 2 h to optimize various physical and chemical properties, including the film lattice orientation, oxygen vacancy concentration, and specific surface area. This approach achieved efficient degradation rates of ofloxacin (OFX, 99.5%), tetracycline (TC, 77.6%), and rhodamine B (RhB, 94.1%), thereby enhancing the photocatalytic degradation efficiency of these pollutants. Additionally, doping with elements such as nitrogen or carbon, or constructing heterojunctions like TiO₂/g-C₃N₄, can broaden the light response range, enhance the charge separation efficiency, and synergistically improve the degradation performance, thereby increasing the overall degradation efficiency. For example, the synthesis of La-doped TiO₂@g-C₃N₄ composite photocatalysts enables g-C₃N₄ to be excited by visible light (Eg = 2.66 eV), resulting in the generation of electrons (e⁻) and holes (h⁺). The electrons then transfer to the conduction band (CB) of TiO₂ through La bridges, while the generated reactive species, including ·O₂⁻, SO₄⁻, •OH, h⁺, and e⁻, collaborate to attack TCH molecules, ultimately mineralizing them into CO₂ and H₂O [118].

3.5.2. Metal–Organic Frameworks (MOFs)

MOFs possess high specific surface areas, tunable pore structures, abundant active sites, and an excellent chemical stability, which provide significant advantages in the adsorption and photocatalytic degradation of pollutants, including antibiotics. The degradation of antibiotics by MOFs primarily occurs through two mechanisms: adsorption and photocatalytic degradation. In terms of adsorption, iron-based MOFs, such as MIL-101 (Fe), acquire a positive charge on their surface when the pH exceeds 7. This property enables them to adsorb negatively charged tetracycline molecules (TC⁻) [119]. ZIF-8@TA adsorbs tetracycline through π–π interactions [120], while hydrophobic MOFs, such as ZIF-8, attract hydrophobic antibiotic molecules via hydrophobic interactions [121]. Additionally, Lewis acid sites in MOFs, such as zirconium–oxygen (Zr-O) clusters, interact with the basic groups of antibiotics, such as amino groups, through acid–base interactions [122]. In the field of photocatalytic degradation, MOFs absorb photons from light sources, including visible light. This absorption causes electrons to transition from the valence band (VB) to the conduction band (CB), resulting in the formation of electron (e⁻)–hole (h⁺) pairs. This process generates reactive species, such as superoxide radicals (·O₂⁻), hydroxyl radicals (•OH), and sulfate radicals (SO₄·⁻), which are produced in the presence of persulfate (PMS/PDS). These reactive species attack antibiotic molecules, disrupting their structures through mechanisms such as ring opening and decarboxylation, ultimately mineralizing them into CO₂ and H₂O [123].

3.5.3. Carbon-Based Single-Atom Photocatalysts

By anchoring metal atoms in a monodisperse form on carbon carriers, distinct active sites are formed. The embedding of metal single atoms can effectively adjust the energy band structure, electronic structure, and surface structure of the light-capturing carrier, thereby modifying the material’s light absorption behavior, charge carrier conversion, and surface adsorption. Therefore, utilizing the characteristics of single-atom catalysts for the photocatalytic removal of organic pollutants from water is an ideal concept [124,125,126]. Liang et al. loaded single-atom barium (Ba) onto g-C3N4 as a carrier and verified its photocatalytic degradation of carbamazepine (CBZ) and diclofenac (DCF), demonstrating that the addition of single-atom Ba significantly enhanced the photocatalytic activity, with the reaction’s first-order kinetic constant being over 70 times that of g-C3N4 [127]. Wang et al. synthesized a hybrid of mesoporous graphitic carbon nitride (mpg-C3N4) modified with single-atom dispersed Ag (Ag/mpg-C3N4) using a co-condensation method, which exhibited a good photocatalytic performance for the degradation of bisphenol A (BPA) as a visible light photocatalyst. The performance improvement may be attributed to the synergistic effect between single-atom Ag and mpg-C3N4 [128]. Yang et al. developed a simple in situ growth strategy to cultivate single-atom cobalt on polymeric carbon nitride, achieving an efficient photocatalytic degradation of tetracycline [129].

New types of catalysts have significant advantages in improving the degradation efficiency of antibiotics, but their application still faces some challenges. For example, there are issues with the recovery and stability of nanocatalysts, the water stability limitations of MOFs, and the high production costs of SACs. Future research needs to further optimize the design and preparation processes of catalysts, as well as assess their long-term performance and ecological safety in real water environments.

4. Hybrid Technology

4.1. Microbial Electrochemical Systems

Biological electrochemical systems (BESs) effectively degrade antibiotics through the metabolic activities of electroactive microorganisms. The primary mechanisms involved are extracellular electron transfer (EET) and enzyme-catalyzed pathways. EET is a crucial process for the transfer of electrons between electroactive microorganisms and electrodes, significantly influencing the efficiency of this electron transfer. EET can be categorized into two modes: direct EET and indirect EET [130]. The direct EET mechanism refers to the process by which microorganisms transfer electrons directly to electrodes via cytochromes located on their cell membranes or through conductive nanowires, such as conductive pili. For example, Geobacter sulfurreducens transfers electrons from inside the cell to the anode through a cytochrome network, oxidizing and degrading tetracycline [131]. At the cathode, Shewanella oneidensis receives electrons through nanowires and reduces the nitro group of chloramphenicol (CAP), achieving a 23% increase in the degradation rate within 24 h [132]. Indirect EET refers to the process where microorganisms secrete or utilize exogenous redox mediators as electron shuttles to accelerate electron transfer. For example, Wang et al. added humic acid (HA) as an electron mediator, increasing the degradation rate of sulfanilamide (SDZ) from 80% to 90% [133]; Cheng et al. used biochar as a conductive carrier to promote a long-range electron transfer between microorganisms, increasing the removal rate of sulfamethoxazole (SMX) from 71.58% to 82.44% [134]. In addition, microorganisms can catalyze the breakdown of specific chemical bonds in antibiotics through functional enzymes, leading to their degradation. For example, laccase can bind to the amino structure of tetracycline (TC) and degrade it into smaller molecules via an oxidative pathway [135]. Similarly, nitroreductase gradually reduces the nitro group (-NO₂) of chloramphenicol (CAP) to an amino group (-NH₂), with the intermediate product nitroaniline (AMCl₂) further undergoing dichlorination [136]. In actual bioelectrochemical systems (BESs), electron transfer (EET) and enzyme catalysis often work in tandem to facilitate antibiotic degradation. For instance, in the BES that degrades chloramphenicol (CAP), a cathodic EET provides electrons, while nitroreductase catalyzes the reduction of the nitro group, ultimately leading to dechlorination [132]. Additionally, the anodic electroactive bacterium Pseudomonas transfers electrons through EET while secreting P450 monooxygenase to oxidize the piperazine ring of ciprofloxacin [137].

Biological electrochemical systems (BESs) can be categorized into two application forms: microbial fuel cells (MFCs) and microbial electrolysis cells (MECs). This technology combines the microbial metabolism with electrochemical redox processes, making it an effective approach for treating difficult-to-degrade antibiotics. The main mechanism of MFCs is the conversion of chemical energy into electrical energy through microbial action, where the microorganisms in MFCs can simultaneously generate electricity and degrade pollutants in wastewater [138]. The MFC consists of a biological anode and a non-biological cathode, where antibiotics serve as electron donors and carbon sources at the biological anode. Exoelectrogenic microorganisms and antibiotic-degrading bacteria attach to the anode, forming a biofilm within extracellular polymeric substances, which is responsible for reducing the overpotential of the parent antibiotics and their metabolites [139]. The MEC builds upon the MFC by adding an external power source, and its mechanism involves obtaining certain chemical products through specific reactions in the presence of a small amount of external energy. Compared to MFCs, MECs can more effectively control the microbial living environment and the electrochemical parameters required for chemical reactions due to the presence of external energy, thereby improving the kinetics and thermodynamics of the reactor. Additionally, continuous electrical stimulation can provide more electrons to the microenvironment and stimulate the microbial metabolism by transferring electrons directly or indirectly to bacterial cells. The stimulated microorganisms rapidly metabolize antibiotics by secreting enzymes [140]. Xue et al. studied the performance of a closed-loop anaerobic bioelectrochemical system in removing sulfamethoxazole. They found that the removal rate of sulfamethoxazole increased from 48.8% on day 80 to 68.6% on day 110. This indicates that the biodegradation efficiency of the antibiotic was significantly enhanced by the electrochemical system. Additionally [141], Jiang et al. reported that photochemical and electrochemical catalytic materials (Fe₀/TiO₂) effectively generated hydroxyl radicals and electrons from the bioanode under visible light, achieving the rapid degradation of tetracycline [142].

However, a single BES still has limitations, such as low electron transfer efficiency, the incomplete degradation of antibiotics, and the risk of spreading antibiotic resistance genes (ARGs); combination technology has become a research hotspot. For example, the removal rate of sulfamethoxazole by the BES-constructed wetland system increased from 71.3% to 82.4%, while also inhibiting the spread of resistance genes [143]; the BES–advanced oxidation system synergistically utilized microbial electricity to drive the Fenton reaction, achieving a tetracycline degradation rate of 90.3% while reducing energy consumption [144]; and in the BES–microalgae system, microalgae assisted the BES through adsorption and photodegradation, achieving a removal rate of over 80% for fluorobenzene within 90 h [145]. In summary, BESs demonstrate great potential in antibiotic degradation, but interdisciplinary collaboration is needed to promote their transition from laboratory to engineering applications.

4.2. Microalgae Technology

The primary mechanisms by which microalgae remove antibiotics include bioadsorption, bioaccumulation, and biodegradation. Bioadsorption refers to the strong adhesion of organic pollutants to the surface of algae, which can subsequently be removed through photolysis [146,147]. The efficiency of bioadsorption in microalgae is highly dependent on the structural characteristics of the target antibiotic, such as its hydrophobicity and the presence of functional groups that facilitate chemical adsorption, as well as the specific species of microalgae acting as adsorbents and the prevailing environmental conditions [148]. Consequently, species with a high affinity for target antibiotics, such as Chlorella and Scenedesmus, can be identified for their potential in antibiotic degradation [149]. Additionally, bioadsorption is influenced by pH; variations in pH can affect the ionization or dissociation of antibiotics in aqueous media, as well as the surface charge of the bioadsorbent, whereby microalgal cells uptake antibiotics, which can bind to intracellular proteins or other compounds. Moreover, the accumulation of antibiotics in microalgal cells may lead to the excessive production of reactive oxygen species (ROS). For instance, microalgae effectively remove sulfamethoxazole and levofloxacin through accumulation followed by intracellular biodegradation [150,151]. Biodegradation has been demonstrated to be the most effective mechanism for the removal of antibiotics mediated by microalgae [152]. This process typically involves both extracellular and intracellular biodegradation, or a combination of the two, where the initial degradation occurs extracellularly, and the resulting degradation products may be further metabolized intracellularly [153]. Algal cells possess complex enzyme systems that catalyze various chemical reactions to degrade antibiotics. For instance, carbohydrate-active enzymes, redox enzymes, and hydroxyl radicals play particularly significant roles in the biodegradation of sulfonamides [154].

To improve the degradation efficiency of antibiotics by microalgae, researchers have proposed various innovative methods, including adaptive cultivation, co-metabolism, and microbial consortium systems. Adaptive cultivation improves the tolerance and degradation capacity of microalgae by pre-exposing them to antibiotic environments. For example, pre-adapted Chlorella increased its removal rate of levofloxacin from 16% to 28. Co-metabolism facilitates the synergistic degradation of antibiotics by introducing exogenous carbon sources. For instance, following the addition of sodium acetate, the removal rate of sulfamethoxazole by Chlorella vulgaris significantly increased from 6.05% to 99.3% [155]. Microbial consortium systems leverage the synergistic effects of microalgae and bacteria to enhance degradation efficiency; for example, a microalgae-activated sludge combined system achieved a removal rate of up to 97.91% for cephalosporin antibiotic [156]. Additionally, immobilized microalgae technology streamlines the harvesting process and improves the tolerance to antibiotics by embedding microalgae in carriers. For instance, Kai-Xuan Huang et al. established a microalgae-biochar composite system, first cultivating Chlorella to the stable phase and then adding biochar, which further increased the removal rate of SMX to 45.7%. This is significantly higher than the removal rate of SMX by Chlorella alone at 34.4% and by biochar alone at 20.0% [157]. These methods provide substantial support for the practical application of microalgae technology.

Despite the excellent performance of microalgae technology in antibiotic degradation, its large-scale application still faces numerous challenges. Firstly, the degradation efficiency of microalgae for antibiotics is significantly influenced by environmental conditions such as pH, temperature, and the properties of the target compounds. For example, the adsorption efficiency of tetracycline in Scenedesmus obliquus varies significantly with changes in pH [158]. Secondly, the complex interaction mechanisms in microalgae–bacteria consortia have not been fully elucidated and require further analysis using multi-omics technologies, including metagenomics and transcriptomics [159]. Additionally, while genetically engineered microalgae can enhance degradation capabilities, their ecological risks, such as the potential spread of resistance genes, must be carefully evaluate [160]. Future research should concentrate on optimizing the design of mixed systems, particularly by improving the antibiotic removal efficiency through innovative combined technologies that integrate microalgae biotechnology, advanced oxidation processes, constructed wetlands, and microbial fuel cells. For instance, Ge and Deng reported the effective removal of enrofloxacin and ciprofloxacin hydrochloride through the application of an Fe (III)-microalgae system [161]. Sun et al. utilized a novel microalgae–bacteria microbial fuel cell (MFC) system to simultaneously degrade fluoroquinolone at the anode, eliminate nitrogen at the cathode, and generate bioelectricity [145].

4.3. Constructed Wetland Technology

Constructed wetland systems are a type of wastewater treatment process that mimics natural wetland ecosystems through artificial construction and controlled operation. These systems primarily rely on the interactions among plants, substrates, and microorganisms to remove pollutants from wastewater through physical, chemical, and biological processes. The effectiveness of the pollutant removal in constructed wetland systems largely depends on the degradation capabilities of microorganisms, the adsorption properties of the substrate, and the absorption and accumulation by plants as auxiliary mechanisms [132,162]. For instance, research conducted by Xian et al. indicates that surface flow constructed wetland systems can achieve a removal rate of over 99% for sulfamethazine [163]. In contrast, research by Hijosa-Valsero et al. shows that the removal rate for oxytetracycline in these systems ranges from 47% to 75% [164]. Additionally, the concentration of antibiotics in wastewater can significantly influence the removal efficiency of antibiotics in constructed wetland systems. A study by Wang et al. found that the presence of antibiotics—specifically ampicillin, tetracycline, bacitracin, and colistin—in water reduced the removal rates of the total organic carbon (TOC), ammonia nitrogen (NH3-N), and nitrate (NO3-) in vertical subsurface flow constructed wetland systems [165]. Due to the limited capacity of traditional constructed wetlands to treat antibiotics—coupled with the fact that the presence of these substances alters the microenvironment within the system, making pollutant removal more challenging—various enhancement methods have emerged to complement constructed wetland technology for antibiotic removal. These methods include the development of electro-enhanced constructed wetland systems and the use of bulrush-based biochar as a substrate to improve the performance of constructed wetlands.

5. Conclusions and Perspective

The widespread use of antibiotics has led to increasingly serious residual issues in the environment, posing potential threats to ecosystems and human health. This article provides a macroperspective on the degradation mechanisms of antibiotics in the natural environment (photolysis, hydrolysis, and biodegradation) and the research progress of various artificial intervention technologies (advanced oxidation processes, physicochemical methods, bioelectrochemical systems, etc.). It summarizes the limitations of natural degradation mechanisms: photolysis and hydrolysis processes are significantly affected by environmental conditions (such as the pH, temperature, and dissolved organic matter (DOM)), resulting in an unstable efficiency and incomplete degradation; biodegradation relies on the diversity of microbial communities, but high concentrations of antibiotics may inhibit microbial activity, and degrading bacteria may become carriers of antibiotic resistance genes (ARGs). Artificial intervention technologies have clear advantages but also face significant challenges: oxidation technologies based on free radicals (such as •OH, SO₄⁻) (e.g., Fenton reaction, photocatalysis) have a high degradation efficiency (80–100%) for antibiotics but may produce toxic intermediates and are limited by pH and coexisting substances; adsorption methods and membrane technologies are easy to operate but face issues such as the difficulty in regenerating adsorbents and membrane fouling; bioelectrochemical systems (BESs) and microalgae technologies have potential for degradation and resource recovery, but scaling up requires the optimization of reactor designs and operating parameters; and new catalysts such as nanomaterials, metal–organic frameworks (MOFs), and single-atom catalysts enhance the degradation efficiency by modulating active sites, but face challenges related to stability, cost, and ecological safety. We created

Table 4. The applicability, limiting factors, and cost levels of various degradation technologies.

Technology	Applicable Environmental Systems	Cost	Limitations
Natural Degradation Methods (Photolysis, Hydrolysis, and Biodegradation)	Surface water, soil	Low, relies on natural conditions, no additional equipment required	Slow degradation rate, limited efficiency, highly influenced by environmental conditions (e.g., light, pH, and temperature). Unable to handle high-concentration or complex pollutants and may generate secondary pollution such as antibiotic resistance genes (ARGs).
Microbial Degradation	Wastewater treatment plants (activated sludge process)	Moderate, requires optimization of reaction conditions for non-active substances	Limited degradation capacity for specific antibiotics; may require acclimation. High risk of ARGs dissemination, requiring subsequent monitoring.
Advanced Oxidation Processes (AOPs)	Industrial wastewater treatment	High, requires oxidants (e.g., H2O2, ozone), catalysts (e.g., TiO2), and energy (UV/electricity)	Low oxidant utilization, potential generation of toxic intermediates, requiring post-treatment. Harsh reaction conditions (e.g., acidic pH), complex equipment.
Supercritical Water Oxidation (SCWO)	High-concentration organic wastewater (pharmaceutical plants)	High operational cost, including equipment, energy, and maintenance costs	Acidic intermediates can severely corrode reactors, necessitating expensive materials. Requires continuous heating to maintain supercritical state, energy consumption higher than conventional biological treatment or AOPs.
Membrane Filtration Technology	Drinking water treatment	High, membrane materials (e.g., graphene and carbon nanotubes) are costly and require regular replacement	Severe membrane fouling issues, frequent cleaning or replacement increases maintenance costs. Concentrate disposal may cause secondary pollution.
Adsorption Method	Emergency or advanced treatment	Relatively high adsorbent cost (e.g., biochar and nano-modified materials)	Adsorbents require regeneration or disposal post-use, regeneration costs are high. Sensitive to dissolved organic matter (DOM) or ionic strength, efficiency easily affected by water quality.
Ionizing Radiation	Medical wastewater, high-risk pollutants	High, equipment costs (e.g., cobalt-60 radiation sources, accelerators) and regular maintenance	High initial investment unaffordable for small-scale plants. DOM and inorganic salts reduce efficiency. Scaling challenges: electron beam suitable for small–medium flow, gamma irradiation is mobile, suitable for fixed installations.
Bioelectrochemical Systems (BESs)	Laboratory or small-scale pilot projects	High, electrode materials (e.g., graphene) and system construction costs	Low electron transfer efficiency, difficult to scale up. Requires continuous power supply (MEC), high energy consumption.
Microalgae Technology	Ecological remediation	Moderate, requires light and nutrients for microalgae cultivation	Antibiotics may inhibit microalgae growth, requiring optimized ecosystems. High costs for harvesting and processing microalgae biomass.
Constructed Wetland Technology	Rural/decentralized wastewater treatment	Moderate, low construction cost but large land footprint	Treatment efficiency affected by seasons and climate, poor performance in winter. Long-term operation may accumulate pollutants, requiring regular maintenance.

to compare the characteristics of various degradation technologies. The actual status of antibiotic degradation technology is objectively introduced, providing readers with a reference guide.

In the future, in terms of design, heterogeneous catalysts should be adopted to optimize the pathways for generating reactive oxygen species and reduce toxic by-products. By combining computational chemistry with experiments, efficient catalysts specific to antibiotic degradation should be screened, and high-efficiency directed catalytic systems should be developed. Research should focus on redox coupling technologies that can inactivate antibiotic resistance genes (ARGs) while degrading antibiotics, exploring microbial community regulation strategies to inhibit the potential spread of ARGs. Strengthening natural–artificial coupling processes, such as constructed wetlands and bioelectrochemical systems, will enhance the long-term treatment capacity for low concentrations of antibiotics. Green degradation technologies based on microalgae or a plant–microbe co-metabolism should be developed to reduce the energy consumption and the risk of secondary pollution. At the same time, in the area of intelligent monitoring and risk assessment, big data and predictive models (such as antibiotic environmental fate models) should be integrated to establish real-time monitoring and risk warning systems, assess the ecological toxicity of degradation by-products, and formulate stricter water quality standards and control policies.

In summary, the management of antibiotic pollution must balance degradation efficiency, ecological safety, and engineering feasibility. Through interdisciplinary innovation, a complete solution from basic research to practical application should be constructed to address this global environmental challenge.