Antibiotics are complex molecular compounds with extraordinary antimicrobial abilities. With the accidental discovery of penicillin back in 1928, an array of antibiotics has been developed afterward. Antibiotics are classified based on mechanism of action, bacterial spectrum, type of activity, and route of administration, but the most useful is chemical structure. However, antibiotics belonging to the same structural class share similar effectiveness and allergic or toxic potential [1, 2]. Antibiotics are being used as a therapeutic drug and have successfully been applied in animal farming as growth promoters and improving feed efficiency [3]. Though its use in feed to improve animal growth is banned in the European Union, its use is common in the USA, Canada, and China [4]. Antibiotics are the most successful drugs in the pharmaceutical industry for human and animal treatment alike. Global antibiotic consumption has increased immensely. An increase of 65% defined daily dose was recorded between 2000 and 2015, leading to an antibiotics consumption rate up to 39% [5]. China uses up to 180,000 tons of antibiotics for both human and agricultural purposes [6].

Tetracyclines (TCs) are the most common antibiotic drugs in the world. This broad-spectrum family of antibiotics is known to inhibit protein synthesis in bacteria besides combat a variety of bacterial infections. TCs are derived from various Streptomyces species, and a total of twenty compounds are introduced in the market as antibiotics [7]. The basic structural element is a tetracyclic ring system with various hydroxyl, methyl, keto, and dimethylamino functional groups [8]. TCs are divided into three classes based on their nature, dosage, and elimination time; oxytetracycline, chlortetracycline, and TC. The current study mainly focuses on TC unless otherwise stated. Structural comparison (Fig. 1) of several tetracyclines and mode of action is beyond the current study’s scope. The readers are referred to the work of Chopra and Roberts [9]. TCs have widespread usage in human therapy, aquaculture, and animal husbandry because TCs have low production prices, high quality, and great purity. TCs is ranked second worldwide in production and usage while first in China [10].

Fig. 1
figure 1

Structural comparison of tetracycline, oxytetracycline, and chlortetracycline

Though tetracycline presents several benefits to human and animal health, its overuse is associated with allergic reactions in humans, bacterial resistance, and significant fluctuations in environmental microflora that detrimental to environmental health. It is linked to the difficult metabolization of TCs in the human and animal digestion system and, therefore, excreted into the environment by human feces and animal excreta up to 50–80% [11]. Hospitals, pharmaceutical industries, and livestock also add the share in the accumulation of TC in the wastewater systems [12]. Domestic wastewater contains low (1 µg L−1), whereas hospital wastewater has a relatively higher TCs concentration, 100 µg L−1 [13, 14]. Figure 2 illustrates the aquatic and terrestrial routes of TCs in the environment. TCs are stable and difficult to oxidize in the environment but are unstable at extreme pHs, forming epi- and anhydro- products supported with their relatively low Henry’s constant (3.45 × 10 − 24 to 3.91 × 10 − 26 atm3 mol−1) that indicate the low volatility, ultimately describing lesser degradability as well. The conventional methods aimed at treating wastewater cannot effectively eliminate large quantities of tetracycline; besides, TCs easily make stable compounds due to their binding potential to Ca+2 and other ions [15, 16]. Therefore, a significant portion is detected in wastewater, surface water, groundwater, sludge, and sediments.

Fig. 2
figure 2

Aquatic and terrestrial routes of tetracycline in the environment

Antibiotics like TC are becoming a serious threat to the environment due to their unaccounted use in subtherapeutic animal growth promotion and human treatment. The environmental fate of antibiotics has emerged as a research topic in the previous decade to address the issue. Long-term detrimental effects of TC would lead to ecological imbalance due to poor degradation of TC in the current treatment processes. It is of utmost importance to counter TCs linked to environmental pollution because it severely impacts human health due to the development of bacterial resistance. Non-efficient wastewater systems are the source of antibiotics in the food web that further ruining human health. Soil seepage is drastically affecting the soil microbial flora. Different removal strategies have been developed to degrade TCs from the environment. However, microbial biodegradation is the future of the removal strategies due to its low capital cost, effectiveness, and process simplicity. The literature is deficient in providing a specific review on tetracycline degradation strategies, proposed degradation pathways of TC in various microorganisms, and the role of resistant genes in TC degradation. Therefore, this review aims to collect data and discuss the merits and demerits of available TC degradation techniques, the impact of anti-resistant genes affecting the microbial biodegradation, new insights into the detection of antibiotics, and limitations and prospects of applying microbial degradation for TC degradation.

Degradation strategies of tetracycline

TCs in the environment can be degraded through several pathways. The degradation mechanism breakdown macromolecules to smaller molecules that are relatively less harmful and non-hazardous. The degradation mechanism produces secondary metabolites that may cause pollution and are challenging to mineralize. The chemical structure of TCs (Fig. 1) is essential to understand its degradation route by different degradation techniques. Several removal strategies for TCs have been developed for TC degradation that can broadly be categorized as non-biodegradable and biodegradable. Each strategy has its own merits and demerits. However, the best strategy would be cost-effective, high degradation efficiency, environmentally benign, and up-scalable to the industry level. Non-biodegradable methods are expensive, energy-intensive, and environmentally less benign. Biodegradation is the cost-effective and most suitable approach to degrade TCs in the soil at the expense of microbial enzyme machinery. Understanding the degradation mechanism of different enzymes at the molecular level is crucial for further scale-up to pilot and industrial scale. However, the microbial degradation pathway of TC by different bacteria is not much explored, except some references for pure cultures are reported [17, 18].

Non-biodegradable route of TC removal

Advanced oxidative processes

The inefficiency of conventional methods of removing TC antibiotics from the wastewater streams resulted in developing advanced treatment technologies, including advanced oxidation processes (AOPs). AOPs are cost-effective as the processes operate with less energy than direct oxidation. The process is based on hydroxyl (•HO) radicals that react with the organic compounds. Hydroxyl radicals (•HO) were reported to have higher oxidative reactivity ( = 2.8 V) than ozone and chlorine gas, but their selectivity is relatively poor [19]. These highly reactive •HO radicals are generated by H2O2 and ozone along with metal or semimetal catalysts. AOPs could produce low toxic intermediates, leading to complete mineralization. Advanced oxidation processes are further divided based on light, catalysts, and ultrasounds such as Fenton processes, ozonation, photocatalysis, UV photolysis, and sonolysis. These processes have been studied for wastewater treatment aiming to degrade micropollutants. Table 1 gives a quick overview of comparison on different non-biodegradable methods regarding advantages and disadvantages, applied for TC degradation.

Table 1 A comparison of non-biodegradable tetracycline methods


Ozone is a well-known oxidizing agent that has been extensively studied for tetracycline degradation. It offers advantages over other technologies, such as no chemical sludge generated after ozonation of organic and inorganic pollutants, ozone decomposes quickly to oxygen in water [20]. Understanding the ozonation mechanism is pivotal in determining its role in TC degradation. Ozonation works by two mechanisms: direct reaction with liquefied ozone or indirect reaction by OH radicals formation [21]. Ozone primarily attacks the double bonds, aromatic ring, and amine functional groups of organic compounds like TC in an aqueous form. Ozone reacts via Direct 1,3-dipolar cycloaddition to C = C, C = N, N = N, and in situ generations of •HO. Highly reactive species are generated during the ozone decomposition, especially at alkaline pH [22]. Ozone reacts with TC in its protonated form [H+] and adds one and two oxygen atoms at C11a-C12 and C2-C3 locations, respectively [23]. C11a-C12 double bond is highly susceptible to ozone attack than C2-C3 double bond because it bears two carbonyl groups while the former contains only one carbonyl group. Unstable epoxide intermediate is generated during this process by the release of an O2 molecule by transient ozonide. Epoxide quickly undergoes a rearrangement with hydroxyl groups at C12, resulting in an oxidized product of size 460 Da. Similarly, another product of size 476 Da was generated at double bond C2-C3 [23].

Several factors, including pH, tetracycline concentration, temperature, ozone generation rate, and catalyst dosage, if applied, play their role in effective TC degradation by ozonation. pH plays a crucial role in the ozonation mechanism. At low pH, the hydroxyl group (OH) at the ring I of TC was non-dissociated while the amide functional group (CO–NH) turns into protonated form. This results in a decrease in electron densities on C2-C3 double bond and amide and keto groups at C1 position via conjugation. The decreased electron densities lower the probability of ozone attack at C2-C3 position. Khan et al. [24] proposed an ozone-dependent TC degradation pathway based on TC chemistry, ozonation mechanisms, and the mass spectroscope (M.S.) spectra of TC products. Ozone complemented TC degradation, at a lower pH of 2.2, resulted in oxidized products of m/z value of 461, 477, 509, at C11a-C12 location, while a product of m/z value of 416 at C2-C3 location. Moreover, ozonation at the said locations gave products of m/z 432, 480, 448, 525, and 496. In comparison to lower pH, higher pH of 7 resulted in only three products. The authors speculated that at higher pH, free radicals non-selectively react with intermediate products of TC leading to low molecular weight final products that need advanced, sophisticated instrumentation to detect and quantify. Gulnaz and Sezer [25] found that 50% of TC was degraded within 10 min of Ozonation at pH 3 while complete removal was seen in 40 min ozonation. The authors further reported ozone efficiency for TC degradation follows the order of pH 3 > pH 7 > pH 11.

In contrast, TC degradation is also reported to be favored by higher pH. Wu et al. [26] studied TC degradation at pH values of 5, 7, and 9 for 5 min of Ozonation. The authors found a direct relation in degradation reaction rate constant with alkaline conditions due to hydroxyl radicals' contribution. Higher pH is responsible for a higher OH radicals concentration that is recognized as more reactive than ozone. The authors concluded a higher rate constant of 0.890 min−1 at pH 9 than pH 5 (0.592 min−1). Ozonation is reported to promote TC degradation but does not lead to its mineralization. Several organic compounds are generated during ozonation, but their mineralization is slow [24].

Catalysts like metal oxides can enhance the ozonation process to remove antibiotics and their products. Catalytic ozonation is categorized as homogeneous and heterogeneous ozonation. Ozone decomposition is favored by transition metal ions in the former category while by solid catalysts in the latter category [27]. A comprehensive mechanism of homogeneous and heterogeneous catalytic ozonation is reviewed elsewhere [27]. Catalyst can easily be separated from the liquid and reuse in heterogeneous ozonation catalysis [20]. ZnO nanoparticles catalyzed ozonation potentially degrade TC due to high surface area and lesser production cost. ZnO-ozonation achieved 94% OTC removal at pH 7 after 5 min at ozone generation rate of 1.38 mg /s. [21]. In another study, 88.52% TC degradation efficiency was achieved by ZnO/γ-Fe2O3 nanocomposite [28]. Cobalt-modified silicate ore (CoSiO) catalyzed ozonation removed 93.2% TC [29]. Another catalyst, cerium-doped compound Mn3Gd7–xCex(SiO4)6O1.5 reported to increase the TC degradation up to 15% [30].

A load of research was published on ozonation and novel catalysts to improve the ozonation efficiency. In a recent study, Ostman et al. [31] reported Sweden’s first full-scale ozonation treatment plant for removing antibiotics. However, the study did not mention about tetracycline removal. Ozonation mechanism still needs to be much explored regarding homogeneous and heterogeneous catalysis for optimum reaction parameters. Catalytic ozonation also faces issues like ozone adsorption on metal surfaces that need to be explored. A negative effect of inorganic ions typically presents in wastewater streams on the ozonation process due to their high affinity towards active catalytic sites should be addressed. Novel stable catalysts with enhanced ozonation efficiency should be developed to lower the ozone dose in the reaction process.

Fenton process

H.J. Fenton first reported the Fenton reaction in 1984 by describing that oxidative potential of H2O2 could be enhanced using Fe2+ as catalyst under acidic conditions [37]. The reactions involved in Fenton processes are presented in Eqs. 11d. Equation 1 is considered the core reaction involved in the Fenton chemistry. A detailed insight into the Fenton and Fenton-like reaction chemistry is already presented in a review by Babuponnusami and Muthukumar [35]. The importance of •OH-related reactions is overwhelming as seen by more than 1700 rate constants reported in the last few decades [38]. An in-depth overview of different Fenton processes, including single-Fenton and coupled-Fenton processes, reaction parameters optimization for wastewater treatment is reviewed by Zhang et al. [39]. Classical Fenton or coupled-Fenton processes have been successfully applied to degrade TC, including photo-Fenton [34], sono-Fenton [40], and electro-Fenton process [41].

$${Fe}^{2+}+ {H}_{2}{O}_{2} \to {Fe}^{3+}+\bullet OH+ {OH}^{-}$$
$$\bullet OH+ {H}_{2}{O}_{2} \to {HO}_{2}+ {H}_{2}O$$
$${Fe}^{2+}+ \bullet \mathrm{OH }\to {Fe}^{3+}+ {OH}^{-}$$
$${Fe}^{3+}+ {HO}_{2}\bullet \to {Fe}^{2+}+ {O}_{2}+ {H}^{+}$$
$$\bullet \mathrm{OH}+ \bullet \mathrm{OH }\to {H}_{2}{O}_{2}$$

Fenton reactions are pH-dependent regardless of the substrate. Optimum pH is reported for Fenton reactions is around 3 [42]. At higher pH, reaction efficiency is lesser due to lower generation of OH radicals, even auto-decomposition of H2O2 is noted [43]. Other influencing factors for classical Fenton reactions are Fe2+ concentration, initial concentration of the target organic compound, and temperature [35]. Fenton and Fenton-like processes have been applied to wastewater treatment plants to degrade anthropogenic compounds such as phenol, nitrophenol, chlorophenol, alkylbenzene sulfonate, etc. [44,45,46,47]. Photo-Fenton processes are mostly reported in the literature for TC degradation. Yamal-Turbay et al. reported total TC remediation and 77% mineralization for 40 mg L−1 TC initial concentration at 71.5 mg L−1 H2O2 and 5 mg L−1 Fe2+ [48]. In another study, TC degradation was observed under black-light and solar irradiation with different concentrations of Fe(NO3)3 and H2O2. Total TC degradation was noted in the presence of Fe(NO3)3 after 1 min black-light irradiation [34].

Electro-Fenton is another environment-friendly process where •HO are generated either indirectly at the cathode or directly by anodic oxidation. The process requires lower pH conditions, i.e., pH 2–4. The following reaction sequence (2–4) occurs at the cathode to generate •HO radicals.

$${O}_{2}+2{H}^{+} \to {H}_{2}{O}_{2} \left(acidic\right)$$
$${O}_{2}+2{H}_{2}O+2e \to {H}_{2}{O}_{2}+2O{H}^{-} \left(neutral, basic\right)$$
$${H}_{2}{O}_{2}+2e \to 2\bullet \mathrm{HO }$$
$${H}_{2}{O}_{2}+ {Fe}^{2+}\to {Fe}^{3+}+O{H}^{-}+ \bullet \mathrm{HO }$$

The process is advantageous in terms of low cost; however, the main disadvantage is the production of Fe(OH)3 sludge that could be avoided using heterogeneous reagents like metal oxides. H2O2 concentration plays a significant role in the degradation of TC, supported by Borghi et al. [49]. The authors conducted a rotational central composite design to investigate influencing parameters temperature, H2O2 concentration, and ferrous ion concentration on residual doxycycline concentration and total organic carbon (TOC) in the system. The authors concluded that only one parameter, H2O2 concentration showed a statistically significant effect with TOC reduction up to 30%. In another study, a comparison of UV irradiation, electro-Fenton, and photo-electro-Fenton processes was carried out. The authors demonstrated photo-electro-Fenton as the best TC degradation method following the order: photo-electro-Fenton > electro-Fenton > UV irradiation. Besides, Fe3O4-graphite cathode was reported to be stable during the degradation process and speculated to be reused, suggesting its valuable potential in TC wastewater treatment [50].

Several researchers studied novel heterogeneous catalysts to avoid Fe(HO)3 sludge production. Zheng et al. [51] prepared a novel Fe@Bacillus subtilis heterogeneous catalyst by impregnating iron (iii) chloride hexahydrate and studied it as Fenton’s reagent with H2O2 addition to degrade TC. The authors reported the complete elimination of TC with negligible iron leaching. Besides, the performance of the catalyst was found to be maintained after three consecutive runs. The ultrasound-assisted Fenton process using magnetite (Fe3O4) catalyst was found to remove 93.6% TC within 60 min of reaction with 31.8% removal efficiency of TOC [52]. Wang et al. reported more than 90.7% TC degradation when authors applied ultrasound to 0.2 mM Fe2+ and 2 mM H2O2 in a novel enhanced sonolysis process [40]. In a recent study, the mesoporous bimetallic Fe/Co catalyst was reported to degrade TC optimally (86% removal rate) at Fe to Co ratio of 2:1, pH 5–9, H2O2 30 mmol, and initial TC concentration of 30 mg L−1 [53]. Iron-loaded granular activated carbon (GAC-Fe) catalyst reported a limited removal of TC 4.94% [54]. The authors speculated that the lower TC removal rates in GAC-Fe are attributed to the limited surface area, poor structure, and functional groups in the two materials.

Though Fenton processes effectively remove recalcitrant organics in wastewater streams, the process has limitations that need to be removed before its application at the industrial level. Iron levels beyond the acceptable limit could result from the Fenton process that would need another expensive unit operation. Also, sludge generation is an oft-cited demerit of the Fenton process. Novel catalysts should be explored, developed, and tested to tackle the abovementioned issues before industrial application.


Photolysis (photodecomposition, photodissociation, or photodegradation) is the degradation of inorganic or organic compounds by solar irradiation. A comprehensive overview of chemical reactions during photolysis is referenced by [55]. It is an important antibiotic degradation way in the natural aquatic environment. Photocatalytic oxidation is a non-toxic, stable, and low-cost method to disintegrate TC. The photolysis behavior of antibiotic in an aquatic environment is reported to be influenced by dissolved organic matter and nitrate because humic acid in the organic matter upon irradiation results in the generation of various reactive oxygen species that influence antibiotic degradation [56, 57]. Photocatalysis of TC is a complex degradation mechanism resulting in numerous polar and non-polar intermediates during the process.

TC photolysis has been studied with several photocatalyst composites, including TiO2, hydroxyapatite nanocomposite [58], zinc oxide–Bi2O3/TiO2 heterostructure [59], CdS–TiO2 heterostructure composite [60], In2O3@ZnFe2O4 heterojunctions [61], Ag–AgBr/AlOOH plasmonic [62], ZnIn2S4@PCN-224 [63], g-C3N4/TiO2/CdS [64], and without catalysts presence [65]. Photocatalyst substantially increased TC degradation rate and overall degradation up to 90% in the case of In2O3@ZnFe2O4 heterojunctions [61]. Zhu et al. identified several intermediate compounds in TiO2-assisted photocatalysis and proposed a TC degradation mechanism [66]. Humic acid may also reduce the photodegradation rate by inner filtering [67]. In addition to solar radiation, TC can also be degraded by UV irradiation [68].

The fate of antibiotics by photochemical is reported to be affected by humic substances [69]. It may enhance photolysis of organic compounds by the generation of reactive oxygen species like hydroxyl radicals, peroxyl radicals, superoxide, and single oxygen [69] or inhibit photolysis by acting as an inner filter and competitively absorb light and photons, thus causing loss to organic matter removal rate [70]. Photocatalyst TiO2 was investigated for photodegradation of TC, emphasizing humic acid concerning reaction kinetics and TC's removal mechanism. The authors found a significant inhibition in TC loss by humic acid due to the surface deactivation of TiO2 and hydroxyl radicals quenching. TiO2-assisted photocatalysis is reported to significantly enhance the TC degradation efficiency in the presence of Cu2+ /Pb2+, SO42−/Cl−1, and humic acid, while the presence of tannic acid, gallic acid, citric acid, Tween 80, salicylic acid significantly decreased the degradation efficiency [71]. Niu et al. observed a slight increase in TC photolysis rate at a low humic acid concentration [72]. It was further confirmed that humic acid could inhibit the oxidation pathways initiated by hydroxyl radicals. Jiao et al. reported that the photolysis of TC followed first-order kinetics, and the rate was dependent on the initial concentration of TC. Reaction rate decreased to 0.0014 min−1 when TC concentration was increased from 10 to 40 mg L−1 [73].

Visible light-driven photocatalysis has gained much attention recently in the remediation of environmental pollution. Novel photocatalysts are being developed to work in visible light spectra [74, 75]. AgI/BiVO4 heterostructured photocatalyst was synthesized in situ that has shown excellent photoactivity for TC decomposition under visible light complemented with structure transformation from double type (AgI/BiVO4) to sandwich-type (AgI/Ag/BiVO4). The authors witnessed 94.91% TC degradation within 60 min while degradation efficiency was much higher than BiVO4 (62.68%) and AgI (75.43%) under similar conditions. Besides, superior mineralization was recorded as 90.46%. Total organic carbon (TOC) was removed within 120 min [76]. The efficiency of a photocatalyst is highly dependent upon the spatial separation of photogenerated electron–hole pair. Therefore, the fabrication of heterogeneous photocatalysts with two different inorganic semiconductors seems an excellent strategy to improve the charge separation of electrons and holes. Luo et al. [77] developed a novel g-C3N4/Bi3TaO7 nanocomposite photocatalyst for superior TC degradation under visible-light spectra. The authors proposed a Z-scheme system reaction mechanism for TC degradation under visible light. In another study, 81.2% TC degradation was reported by BOC/BWO Z-scheme photocatalyst [78].

In brief, photolysis is an emerging technology with potential applications to treat wastewater at an industrial scale. Novel photocatalysts are being developed to improve the degradation efficiency and encounter the problems like visible light adsorption, fast electron–hole recombination, and bandgap energy. Photolysis for TC degradation is tested on the laboratory scale. Before its industrial application, the technology needs to be tested at a pilot scale to generate techno-economic data for the stakeholders.

Biodegradable route of TC removal

Microbial degradation of TC

Microorganisms could initiate the biodegradation of toxic compounds like TC by opening their loop structure or cutting the attached functional groups. However, a limited number of pure microbial strains are isolated with TC degrading potential (Table 2). Several putative biodegradation pathways of TC have been proposed by different researchers based on biodegradation metabolite products. Yin et al. [18] proposed three putative TC degradation pathways by Klebsiella sp. strain TR5 based on 8 metabolites: i) reduction of a hydroxyl group on C-3 of TC followed by successive dehydration reactions at C-12-a and C-6; ii) demethylation on C-4; iii) oxidization of TC on C-5 followed by removal of carbonyl group on C-1 (Fig. 3). In another study, Klebsiella sp. strain SQY5 is reported to start TC biodegradation by removing the methyl functional group. Once the hydrolysis opened the ring, the carbonyl group was removed, followed by removing the amine group that leads to successive removal of further two methyl and three hydroxyl groups. During this degradation process, the authors identified nine degradation products and proposed a new putative degradation mechanism [79]. Leng et al. [17] reported that bacterial strain Stenotrophomonas maltophilia DT1 initiates TC degradation by demethylation at C-4, followed by carbonyl and amine group removal. The authors identified six biotransformation products during TC degradation. In all the studies reported, it is common that N-methyl, carbonyl, and amine groups are removed from the parent compound during TC biodegradation.

Table 2 Pure microbial strains degrading tetracycline
Fig. 3
figure 3

TC biodegradation pathway proposed in Klebsiella sp. strain TR5 [18], reprinted with permission from Elsevier. Copyright [2020]

Liao et al. [83] reported microbial degradation of chlortetracycline (CTC) by a mixed microbial community with and without acclimation. The acclimated microbial community showed more effective degradation efficiency. An increase in acclimated microbiota concentration increased biodegradation. Acclimated microbial community resulted in 48.7% and 84.9% CTC removal rate in 1 and 4 weeks, respectively, compared to 15.9% and 32.5% for the non-acclimated microbial community. The authors further reported a higher 89.8% removal rate at 45 °C while only 18.8% at 5 °C.

TC is found in reasonable quantities in wastewater streams, as stated earlier in this text. Many researchers have studied fate and TC’s occurrence in activated sludge systems [16, 84]. TC has been reported non-biodegradable in sludge systems because it can easily sorb up to 90% of sludge by binding to Ca+2 and similar ions to form stable complexes [85]. Cetecioglu et al. [86] reported that > 80% TC introduced in the anaerobic reactor was fully or partially degraded. However, complete mineralization was not speculated. TC biodegradation showed a non-biodegradable behavior under nitrate-reducing and sulfate-reducing conditions, while a slightly biodegradable behavior was seen under methanogenic conditions [87]. In contrast, Klebsiella sp. SQY5 showed a maximum TC reduction rate and denitrification of 0.113 mg L−1 h−1 and 4.64 mg L−1 h−1, respectively, after 32 and 92 h of inoculation.

Activated sludge is the prime source for novel microbial strain exploration. In a recent study, Yang et al. [88] explored sludge for TC degrading microbial strains and reported four antibiotic-degrading bacterial strain roots to Pseudomonas, Bacillus, and Clostridium genera. Pseudomonas strains showed increased biodegradation potential under aerobic while Bacillus and Clostridium strains under anaerobic conditions. The authors further suggested that sludge contains 24 bacterial genera with potential antibiotic-degrading capability.

Soil is the main harbor of microorganisms. Bacterial community has shown tolerance towards TC, OTC, and CTC and co-tolerance towards Cu indicating the metabolic potential of bacteria to further utilize these compounds leading to ultimate degradation [89]. Therefore, TC-contaminated soil could be a potential source for isolating bacterial strains capable of degrading TC, possibly employing novel degradation and bio-geo-chemical pathways. Wu et al. [90] isolated Raoultella sp. XY-1 and Pandoraea sp. XY-2 from TC-contaminated soil capable of degrading 81.72% TC within 12 days in Lysogeny broth medium. The evaluation of the microbial community structure of tested soil during TC degradation revealed four predominant phyla, including Proteobacteria, Bacteroidetes, Acidobacteria, and Chloroflexi. The nutrients available in the soil also affect the biotransformation potential of microorganisms. Stenotrophomonas maltophilia strain DT1 was tested against various nutrients, i.e., No background nutrient (N.B.), peptone (P), peptone plus citrate (P.C.), and peptone plus glucose (P.G.) for TC biotransformation. The strain showed the order of PC > P > PG > NB = 0 for biotransformation.

Membrane bioreactors (MBR) have been successfully tested for treating wastewater and are considered superior over conventional wastewater treatment. The impact of TC on semi-industrial MBR was tested for denitrification, coupled with TC biodegradation. No disturbance in MBR was recorded on the injection of TC to eliminate organic matter and nitrification. However, denitrification was affected slightly. It was further confirmed that the origin of activated sludge has a positive impact on MBR performance for TC removal. Activated sludge from a conventional reactor can withstand TC concentration up to 10–20 mg TOC L−1 [91].

In comparison, sludge from the MBR reactor can operate optimally up to TC concentration of 40 mg TOC L−1 with 90% of TC removal using acclimated sludge. The authors also verified a robust inhibitory effect on the microbial community by TC, but phylum Proteobacteria showed resistance up to 1–1000 µg L−1 [92]. Likewise, Tran et al. [93] showed a higher TC removal (83.3–95.5%) by MBR than the conventional process (44.3–87.6%). Sheng et al. [94] observed 100% TC removal for lower (≤ 1 mg L−1) TC concentration in MBR utilizing nitrosifying sludge but low removal (40%) at higher (10 mg L−1) concentration. The data indicated that MBR is a better option than conventional wastewater treatment plants for TC removal. It can further be affirmed that MBR can effectively remove TC in wastewater streams when in a lower concentration. However, to remove higher TC concentration, integration with other removal strategies like Ozonation would be helpful.

TC biodegradation is also studied in integration with other removal strategies. Simultaneous sorption and TC biodegradation were studied by nitrifying granular sludge [95]. The pseudo-second-order kinetic model was verified by the adsorption process, in addition to the complex mechanism of surface adsorption and intra-particle diffusion process. A successive increase in TC removal rate was noted up to 7.30 mg g−1 with an increase in TC initial concentration up to 30 mg L−1 with non-sterilized granules, but the maximum percent degradation verified was 51.76% at an initial concentration of 20 mg L−1. The authors explained that the sorption of TC to nitrifying granular sludge occurred quickly and reached the maximum sorption capacity after 4 h. At the same time, biodegradation proceeds gradually and continues till 50 h of operation. However, Kim et al. [96] found no evidence for TC biodegradation in activated sludge and reported sorption as the principal TC removal mechanism.

Microbial fuel cell (MFC) is another approach in the primitive stage to biodegrade TC. MFCs offer the advantage of utilizing organic compounds from readily biodegradable to refractory organics and convert them into electricity. Electrodes in MFCs act as continuous electron acceptors for anaerobic microorganisms, resulting in enhanced organic matter degradation under anoxic conditions. Wang et al. [97] studied a novel approach of anaerobic degradation of TC by MFC using glucose-TC mixtures under gradient acclimation conditions. The authors reported 79.1% TC degradation within 7 days of operation, much higher than the conventional anaerobic method (14.9%). In another study, Wang et al. [98] reported 74.2% and 78% degradation rate of CTC and OTC, respectively, by MFC. Since MFCs could be open circuit and closed circuit, it was found that TC removal efficiency is higher in closed circuits MFCs than open circuit ones.

An MFC coupled constructed wetland was successfully tested for wastewater containing TC and sulfamethoxazole. Zhang et al. [99] reported significant decrease in TC and sulfamethoxazole concentration with the development of resistance genes. Another report found that the 3D-biofilm electrode reactor removed TC up to 95% [100]. Since TC degradation by MFCs is at a developmental stage, limited data is available in the literature that also points out research potential in this research area.

Factors affecting microbial TC biodegradation

Microorganisms thrive best at specific parameters for optimum growth and metabolic activities like temperature, pH, and nutrient conditions. To optimize the TC biodegradation by various bacteria, researchers have optimized the reaction conditions, including pH, initial TC concentration, and temperature. Leng et al. [17] studied Stenotrophomonas maltophilia DT1 at different pH conditions for optimum TC biodegradation. The authors verified the fastest TC hydrolysis rate at initial pH of 10; the hydrolysis rate was increased with an increase in pH. The highest TC biotransformation was visualized at pH 9. In comparison, a lag phase of 3 days was observed when pH was 6 at the reaction initiation.

The initial concentration of antibiotics like TC is vital in determining biodegradation potential and the rate of biodegradation by microbial communities. Degradation ratio of Klebsiella sp. SQY5 is reported to increase with an increase in initial TC concentration when tested from 10 to 100 mg/L. The maximum degradation ratio reached up to 89.66% for 80 mg/L initial TC concentration. However, the degradation ratio tends to decrease afterward, but it is interesting to know that since strain SQY5 can utilize TC as a carbon and energy source, it can withstand a selection pressure of 100 mg/L TC but with a lower degradation ratio [79].

Stenotrophomonas maltophilia DT1 is reported to show the Michaelis–Menten model of biotransformation kinetics of TC as a function of initial TC concentration. The degradation rate of DT1 strain increased to 75 mg/L initial concentration; afterward, it decreased [17]. Klebsiella pneumonia was identified from chicken manure that can degrade up to 90% TC within 36 h at an initial concentration of 200 mg/L under optimized conditions [18]. Sphingobacterium sp. PM2‐P1‐29 reported degrading TC with up to ~ 50% decrease in OD363 for a 27-h incubation period [80].

In addition to bacteria, yeast strain Trichosporon mycotoxinivorans XPY-10 was studied for TC biodegradation at a concentration of 600 mg/L. The strain managed to degrade TC up to 78% within 7 days [82]. Recombinant E. coli strain ETD-1 was reported to degrade 95% doxycycline within 48 h at an initial concentration of 50 mg L−1 [101].

As far as temperature is concerned, little info is available on this aspect. Shen et al. [102] noted that TC degradation rates increased in swine manure by an increment in temperature up to 55 °C where maximum TC degradation was recorded. During composting, thermal degradation and biodegradation work synergistically to improve TC degradation [103]. Ratasuk et al. [104] investigated the role of temperature, illumination, and pH on the degradation of OTC in swine manure and noted that 65% OTC was degraded after 100 h at 40 °C. Li et al. [105] also stated that temperature and pH are major factors degrading OTC. Limited literature is available on pure microbial strains able to degrade TC. Therefore, future research must be directed to isolate novel bacterial and/or fungal strains for TC degradation.

Enzymatic degradation of TC

Enzymatic preparation has been effectively utilized in removing TC from the environment. Laccase has been extensively used to catalyze the oxidation of a wide range of phenolic and non-phenolic compounds, including environmental pollutants. Therefore, laccase can offer a green alternative in bioremediation and wastewater treatment [106]. Laccases are the largest subgroup of multicopper oxidases widely found in bacteria, fungi, and insects [107]. Laccases are dimeric or tetrameric glycoproteins containing four Cu atoms organized into an active site [108], serving as a backbone for the catalysis process of various substrates, including tetracycline.

The laccase catalysis mechanism involves three steps: a) reduction of type I Cu; b) flow of electrons from type I Cu to type II and type III cluster; c) oxygen reduction to water [109]. The overall mechanism involves four-electron O2 reduction to water. Initially, the enzyme is completely oxidized at the resting stage. Then at the T1 site, Cu centers reduction (Cu2+ to Cu+) occurs by the substrate. After that, dioxygen reduction is initiated at T2/T3 cluster site, forming peroxide intermediates and their subsequent reduction to water molecules. In the final step, all the four Cu centers are oxidized with the release of O2− as a second H2O molecule. The net result is the oxidation of four substrate molecules to create four radicals while reducing one oxygen molecule to two water molecules [106]. In another study, Suda et al. [110] proposed that TCs biodegradation mechanism mainly involves the oligomerization of oxidized TCs via radical–radical coupling by laccase in the presence of 1-hydroxybenzotriazole (HBT). Crude lignin peroxidase from Phanerochaete chrysosporium degraded 95% of 50 mg L−1 TC and OTC at 40 U L−1 enzyme activity [111]

Enzyme immobilization techniques have been developed to improve enzyme stability, reusability, and lower the cost. Enzymatic membrane reactors have been developed based on laccase immobilized on ceramic membranes to degrade TC in wastewaters. Cazes et al. [112] reported 56% TC degradation with an enzymatic membrane compared to 30% for free laccase. In another report, laccase from Trametes versicolor was studied in relation to the pore diameter of the membrane and gelatin concentration for TC degradation. The authors reported a higher specific degradation rate of 175 mg h−1 m−2 at enzyme and gelatin concentration of 10 g L−1 and pore diameter of 1.4 µm [113]. The data indicated that improved reactivity and stability of immobilized enzymes have a positive impact on TC degradation.

Magnetic cross-linked enzyme aggregates (M-CLEAs) are a relatively new immobilization method where CLEAs are attached to amino-functionalized magnetic nanoparticles. Characteristic features of CLEAs are presented in the work of Sheldon [114]. To lower the cost of operation, M-CLEAs are utilized, which present easy separation and recycling of immobilized enzymes from the reaction mixture by the magnetic field [115]. Yang et al. [116] studied M-CLEAs prepared from Cerrena laccase for antibiotic treatment and observed effective degradation efficiency for TC, oxytetracycline (OTC) followed by ampicillin, sulfamethoxazole, and erythromycin. Cerrena laccase eliminated 100 µg ml−1 TC in 48 h at 25 °C. Though it is reported that T. versicolor laccase removed 78% of 100 mg L−1 TC in 18 h in the absence of a mediator [117], Cerrena laccase is proved to be more effective in TC degradation [116]. It is worth mentioning that the redox mediator ABTS was unable to improve the degradation of TC and OTC, although it is the best mediator for the decolorization of Coomassie Brilliant Blue by C. laccase.

Laccase oxidation system was coupled with soil adsorption to study the simultaneous removal and biodegradation of multiple antibiotics [118]. Laccase-hydroxybenzotriazole and laccase-syringaldehyde showed higher TC, OTC, CTC, and DC. (doxycycline) removal, while the only laccase showed poor removal efficiency comparable to the control reactor (no laccase). Laccase-mediated oxidation systems and soil adsorption systems resulted in 70% antibiotics removal in just 15 min of reaction while close to 100% in 180 min. In another study, Migliore et al. [119] reported the complete removal of OTC by extracellular laccase by fungus Pleurotus ostreatus in 14 days. Likewise, crude lignin peroxidase (40 U L−1) was reported to degrade 95% of TC and OTC at an initial concentration of 50 mg L−1 in about 5 min.

Although enzyme-mediated TCs biodegradation is a practical approach with promising results, future research should focus on developing cost-effective sorbent materials, innovative immobilization techniques for multifunctional catalysts with higher catalytic potential. Novel enzymatic bioreactor configuration must be sought for higher degradability with minimum input on enzyme concentrations. Most of the research has been conducted at a lab scale. Therefore, enzyme-mediated degradation must be tested under real field conditions and scale-up to pilot scale. In this regard, bioreactor design, i.e., packed bed or membrane-based for continuous operation, must be carefully decided for optimum results.

Antibiotic resistance genes (ARGs) relation to tetracycline degradation

Antibiotics resistance refers to an increase in minimum inhibitory concentration (MIC) of antibiotics towards microorganisms. In this mechanism, microbes survive antibiotics stress [120]. ARGs are natural in the environment, but their increased prevalence by human activities has led them to consider ARGs as an emerging environmental contaminant [121]. ARGs transfer from environmental bacteria to human pathogens resulting in decreased antibiotic treatment efficiency, thereby causing severe health concerns. Global tetracycline-resistance was 8.7% and 24.3%, respectively, for methicillin-resistant Staphylococcus aureus and Streptococcus pneumoniae [122]. Tetracycline resistance (TC-resistance) is acquired in the following ways: acquiring some genetic material already carrying resistance genes, mutations at ribosomal binding sites, and chromosomal mutations resulting in overexpression of intrinsic resistance [123]. Resistance to TC is mainly conferred by 1 or more than 36 tet genes [124].

Molecular TC-resistance mechanisms include efflux, ribosome protection proteins, and enzymatic inactivation [125]. Seven efflux pump groups are identified, so far, conferring TC-resistance, most members lie into facilitator superfamily. Only three genes have been reported to inactivate TC [126], while only one enzyme Tet(X) demonstrated confirmed activity in vitro. Tet(X), a flavoprotein monooxygenase, utilizes monohydroxylation to inactivate TC followed by the non-enzymatic breakdown [127]. Nesme et al. [128] reported that tetracycline inactivation is rarely detected in environmental metagenomes. Monooxygenases like Tet(X) are commonly found in nature, catalyzing a range of chemical transformations, including aromatic modifications. Therefore, sequence diversity of flavoenzymes could result in the acquisition of new functions by horizontal gene transfer. So far, limited information is available in the literature on Tet(X) for environmental samples regarding TC biodegradation that warrants future research in this area because enzyme might have been under sampled and could be a potential source of undiscovered inactivating enzymes [129].

Novel antibiotics detection techniques

The conventional antibiotic detection methods are slow and do not provide quantitative data on the antibiotic residues in the environmental samples. Therefore, it is necessary to develop novel antibiotic detection techniques that are rapid, robust, cost-effective, time-saving, and easy to perform. Antibiotic residues in environmental samples are currently detected by high-performance liquid chromatography, gas chromatography–mass spectrometry, and liquid chromatography–mass spectrometry. Chromatography is the most reliable antibiotic detection technique, but it requires expensive instruments, trained personnel, and laborious pretreatment of samples.

Biosensors, in contrast, can circumvent these challenges for rapid and accurate on-site analysis. Biosensors are compact analytical devices with only two main components, a) target signal element, b) signal transducer element. Various biosensors have been developed, including electrochemical, mass-based, optical, enzyme-based, immunosensors, microbial biosensors, aptasensors, calorimetric, and molecularly imprinted polymer sensors. Biosensors selection criteria are based on the intended application. Each biosensor has its potential in a specific field with limitations. The working principle of these biosensors is beyond the scope of the review. The readers are referred to [130, 131]. Biosensors offer several advantages: high selectivity detection, high sensitivity, rapid identification, lesser inhibition to the solute, repeated biocatalyst usage, and applicability to a range of solutes from colorless to colored/turbid samples, portability, on-site pollutants detection, and detection of the targeted compound in complex matrices. Various biosensors have been studied for TC detection, including aptasensors [132,133,134], calorimetric [135, 136], electrochemical [137] immunosensor [138,139,140]. For an in-depth understanding of electrochemical and optical aptamer-based sensors, the readers are referred to Jalaian et al. [141]. Delgado et al. [142] developed an ultralow cost electrochemical sensor made of potato starch to detect trace amounts of TC (detection limit–1.15 µmol L−1) in drinking water.

With the advancement in nanotechnology, new sensors are being developed. Nanofabrication on biosensors showed substantial advantages; a) improved optical and electrochemical measurement, b) high portability, and more practical applications. Zhang et al. [143] reported an aptamer nano-biosensor based on nanoporous silicon for the rapid detection of tetracycline. Electrochemical impedance spectroscopy determined a decrease in biosensor impedance with the specific binding of TC. The linear range reported for the sensor was 2.1–62.4 nM. A highly efficient nano-biochemical sensor was developed in a recent report to sense the ultralow concentration of TC (10–9 mol L−1) using hot spot surface-enhanced Raman scattering (SERS). The sensor showed TC detection with an enhancement factor of 2 × 108. It was verified that the sensor works optimally at a pH range of 5 and 6 [144] that presents their limitation under harsh conditions.

Biosensors have been developed robustly in the current decade. Real-time, high sensitivity, selectivity, and easy-to-handle detection devices in environmental pollution detection and clinical diagnostics require ever-increasing demand for biosensors development. Much advances in biosensors for biomedical applications are sought, but, in comparison, environmental monitoring is still at the infancy stage due to limitations of inherent challenges in the environmental analysis [145]. In the future, biosensor technology should be focused on the development of microarray and nano-biosensors for multiplexed analyte detection for optimum results under field conditions. Electrochemical and enzymatic biosensors have been extensively used in environmental monitoring, but these biosensors work under a narrow range of optimum temperature and pH. Therefore, research efforts must be put in search of stable materials to improve the biosensor's detection capability under harsh conditions.

Limitation and prospects for bioremediation of tetracycline

Antibiotic contamination is a global threat that has increased the risk of antibiotic resistance in microorganisms. Following the limitations and proposed prospects for future research concerning the degradation of TC: A) conventional wastewater treatment plants should be upgraded with novel bioremediation techniques to treat higher concentrations of tetracycline in wastewater streams effectively; B) most of the research has been conducted on a laboratory scale. Therefore, it is necessary to put efforts towards pilot-scale or industrial-scale research to generate data for policymakers; C) TC degradation occasionally results in complete mineralization due to the toxicity of intermediate products and/or byproducts of the TC degradation pathway that needs more research in this area; D) single degradation processes are not viable for the complete removal of TC from the aqueous and terrestrial environment. Therefore, future research must be directed towards the integration of different degradation pathways like ozonation and adsorption, photocatalysis and ozonation, and most importantly, ozonation and biodegradation; E) activated sludge and contaminated soil have not been fully explored for their potential with respect to the discovery of novel degrading bacterial strains. Much research is needed to discover novel bacterial strains; F) future research should be focused on Omics technologies and recombineering to isolate genes responsible for TC degradation and the development of mutant strains with dedicated antibiotic degradation capacity; G) new cheap materials should be discovered for adsorption processes and biosensor fabrication to reduce the cost; H) novel modification processes should be sought to improve the oxidation at the anode in electrochemical oxidation during the photocatalysis process; I) new nano-sized photocatalysts should be developed that can work effectively in visible light.


Tetracycline antibiotic has widespread usage due to its low cost, effectiveness against bacterial infections, and as a growth promoter in livestock. However, a higher proportion of tetracycline is recovered in the waste stream due to lower biodegradability in human and livestock and poses a severe threat to environmental ecology and human health. Complete tetracycline mineralization is not possible in currently available biodegradable and non-biodegradable routes due to the formation of inhibitory intermediate products and byproducts during the degradation process. The identification of intermediate products is mandatory to propose the degradation pathways. Integration of ozonation and microbial biodegradation is the recommended process for effective removal of tetracycline from aqueous streams. Initial tetracycline concentration, pH, and temperature are found to be the critical factor affecting tetracycline biodegradation. The current study systematically compiled TC degradation strategies and identified pros and cons in each strategy. The study finally concluded with future perspectives and pointers for the decision-makers to work in association with the stakeholders to tackle tetracycline degradation, thereby providing a safer environment to live in.