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Article

Degradation of Antibiotics in Aquaculture Seawater: A Treatment Based on Ozone Assisted with Hydrodynamic Cavitation

by
Xiaodian Huang
1,2,*,
Dong Yang
1,
Liang Song
1 and
Yongcan Jiang
1
1
Power China Huadong Engineering Corporation Ltd., Hangzhou 311122, China
2
College of the Environment and Ecology, Xiamen University, Xiamen 361102, China
*
Author to whom correspondence should be addressed.
Water 2025, 17(4), 566; https://doi.org/10.3390/w17040566
Submission received: 13 January 2025 / Revised: 12 February 2025 / Accepted: 13 February 2025 / Published: 15 February 2025
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Abstract

:
Antibiotics in aquaculture pose significant environmental risks due to their widespread distribution in water, impacting ecosystem health. To address this hot issue, an ozone-assisted hydrodynamic cavitation (OAHC) system was developed for the efficient treatment of aquaculture seawater contaminated with antibiotics. The system demonstrated remarkable efficiency, achieving complete degradation of eight antibiotics within a reaction time of 20 s. At the same time, water quality parameters, such as dissolved oxygen (increased from 9.79 mg/L to 13.19 mg/L) and nitrite nitrogen (reduced from 0.14 mg/L to 0.01 mg/L), significantly improved post-treatment. The OAHC-based system minimized harmful by-products, ensuring compliance with Chinese water quality standards. As a supplementary study, a laboratory-based simulated experiment was conducted with FLO as the target antibiotic. The investigation of kinetics and mechanisms indicated that •OH plays a predominant role in the OAHC-based aquaculture seawater treatment system. As global regulations tighten on antibiotic discharge, OAHC-based technology is poised to become a cornerstone of next-generation water treatment solutions. Future research should prioritize field-scale validation and real-time monitoring to accelerate industrial adoption.

1. Introduction

Aquaculture is predicted to play a critical role in providing nutritious food to satisfy the rising global demand for seafood and omega-3 fatty acids [1]. According to the statistics of the Food and Agriculture Organization of the United Nations (FAO) [2], global aquaculture production reached 130.9 million tonnes in 2022, valued at USD 312.8 billion, accounting for 59 percent of global fisheries and aquaculture production. Specifically, marine and coastal aquaculture contributed 37.4 percent to the production volume of farmed aquatic animals. The rapid growth of aquaculture, which is mainly attributed to the expansion and intensification of aquaculture activities, has also led to an increased risk of pathogens and infectious diseases [3]. In this context, antibiotics use may serve to increase growth and substitute for good animal husbandry practices [4].
Antibiotics, also known as antibacterial agents, are a class of substances with the ability to kill or inhibit the growth of microorganisms. These substances have been widely utilized in both human and veterinary medicine for the prevention and treatment of bacterial infections [5]. However, a significant proportion of antibiotics consumed by humans and animals remain unmetabolized, with more than half being expelled into the environment via urine and feces [6]. The presence of antibiotics in the environment imposes selective pressure on microbial populations, thereby promoting the prevalence of antibiotic-resistant genes (ARGs), which represent a substantial threat to public health [7]. According to previous research, incomplete metabolized antibiotics and ARGs have been widely detected in surface water, including rivers [8,9,10], lakes [11], groundwater [12,13], seawater [14,15,16], and even drinking water [17,18].
Compared with antimicrobial use in terrestrial food animal production, the application of antibiotics in aquaculture provides a potentially wider environmental exposure pathway for drug distribution through water, with important ecosystem health implications [4]. Aquaculture effluent has been regarded as an important pollution source of antibiotics and ARGs in coastal environments [14]. In coastal mariculture sites, the concentrations of antibiotics are markedly elevated compared to other locations [19]. Furthermore, the coastal zone, as the interface between terrestrial and marine environments, is a major repository for microplastic pollutants, which enhance the dispersal risk of ARGs [16]. Disinfection and oxidation processes, such as chlorination, ozonation, and UV irradiation, have been widely used for water and wastewater treatment to control pathogens and chemical pollutants [20]. However, chlorine and ozone are not effective in completely eliminating certain antibiotics, such as florfenicol (FLO), while UV irradiation requires an excessively high light intensity, which poses significant challenges for implementation in flow-through aquaculture systems due to their high flow rates [21]. Consequently, it is imperative to develop reliable and effective technologies for the removal of antibiotics from aquaculture wastewater.
Aquaculture water treatment typically encompasses either a singular process or an integrated approach combining physical, chemical, and biological methodologies. However, conventional physical and biological methods exhibit limited efficacy in removing antibiotics. Consequently, chemical methods are predominantly utilized for the removal of antibiotics from water [22]. Among chemical methods, advanced oxidation processes (AOPs) have become an important area of study worldwide in the treatment of organic pollutants due to their high efficiency, wide application range, and thorough reaction in treating refractory organic pollutants [23].
The principle of AOPs is to generate free radicals, such as hydroxyl radical (•OH), through the action of light, electricity, catalysts, etc., which can react with pollutant molecules. Among the AOPs, ozone-based advanced oxidation processes (O3-AOPs) have emerged as one of the most competitive and appealing wastewater treatment technologies due to the strong oxidation and relative environmental friendliness of ozone [24]. There are several methods to enhance the decomposition of ozone to yield •OH radical, including exposure to ultraviolet (UV) irradiation or the addition of H2O2. Liu et al. [25] introduced ozone gas at a rate of 30 mg/min into ciprofloxacin solution coupled with UV irradiation (with light intensity stabilization at 2.60 ± 0.03 mW/cm2), achieving complete degradation of 100 mg/L ciprofloxacin within 60 min. Chen and Wang [26] introduced ozone gas into 40 mg/L hydrogen peroxide solution to generate •OH, which completely degraded 20 mg/L of ofloxacin within 5 min. However, due to the relatively low concentration of ozone at the initial stage of the reaction, excess H2O2 may act as a •OH scavenger, leading to the inhibition of ofloxacin degradation. These processes are rather costly and energy-intensive [27], which restricts their large-scale industrial applications.
In recent years, micro-/nano-bubbles, characterized by diameters less than 100 μm, have garnered considerable attention due to their distinctive physicochemical properties. The combination of ozone and micro-/nano-bubbles has been widely adopted across various sectors, particularly in environmental engineering, owing to their exceptional ability to enhance gas solubility and generate hydroxyl radicals [28]. These bubbles are generated through various methods, such as pressure saturation, mechanical agitation, bubble shear, and splitting, which are based on the hydrodynamic cavitation mechanism of liquid and gas flow. Among these techniques, venturi-type generators have numerous advantages, such as (i) low pump power, (ii) compact size, and (iii) high-density generation of micro-/nano-bubbles [29].
In previous studies, an ozone generator based on strong ionization discharge in combination with the micro-streamer and micro-glow discharges was developed. This generator, when coupled with a Venturi mixer, has been successfully utilized for the treatment of ballast water [30] and drinking water [31]. However, it remains uncertain whether this technology can be effectively adapted for application in the more complex domain of aquaculture water treatment. In this paper, an ozone-assisted hydrodynamic cavitation (OAHC) system was developed for the efficient treatment of aquaculture seawater contaminated with antibiotics. The antibiotics utilized in this study encompass three distinct classes, as detailed in Table 1. Firstly, the degradation of antibiotics during the aquaculture seawater treatment process was systematically assessed using the OAHC-based system. At the same time, water quality data were collected to assess the overall performance of the OAHC-based system. Subsequently, FLO was selected as the target antibiotic to study the kinetics and mechanisms of antibiotic degradation in the OAHC-based system.

2. Materials and Methods

2.1. Chemicals

The antibiotics (listed in Table 1) used in this study were purchased from J&K Scientific (Beijing, China) with a purity exceeding 97% and were utilized without any additional purification. All other chemicals, excluding the solvent used for LC-MS, were all of analytical grade. All aqueous solutions were prepared with deionized water obtained from a Milli-Q System (Millipore Corporation, Billerica, MA, USA). Solid-phase extraction (SPE) cartridges (Oasis-HLB®) with polymeric reversed stationary phase were purchased from Waters (Milford, MA, USA).

2.2. Experimental Setup and Procedures

2.2.1. Aquaculture Seawater Treatment

The simulated experiment for the treatment of aquaculture seawater was conducted at Xiao Deng Island, Xiamen, Fujian province, China. A sketch of the system for aquaculture seawater treatment is shown in Figure 1. Sand-filtered seawater was continuously supplied to each aquaculture tank for Epinephelus coioides culturing. The aquaculture seawater, flowing at a rate of 10 m3/h, underwent pretreatment through filtration, foam flotation, and microbiological processes. The pre-prepared high-concentration antibiotic stock solution was infused into the pipeline via a peristaltic pump, resulting in an aquaculture seawater solution with a concentration of 100 μg/L for each individual antibiotic. Subsequently, the solution containing reactive oxygen species (ROS) was introduced into the pipeline at a flow rate of 1 m3/h through a jet device. The samples were collected in bottles containing sodium thiosulfate from a designated sampling site, corresponding to a reaction time of 20 s. Prior to analysis, the collected samples were subjected to SPE for desalination and enrichment.
The ROS generator, based on ozone-assisted hydrodynamic cavitation (OAHC), which was used in this study, primarily consisted of an ozone generator [32] and a venturi mixer (Mazzei 0287, Mazzei Injector Company, Bakersfield, CA, USA). The ozone generator produced oxygen activated species (OAS), primarily consisting of O2+, O(1D), O, O2−, O2(a1△g), and ozone, with an average concentration of 118 ppm, as measured by an online monitor (BMT 964C, BMT Messtechnik GmbH, Berlin, Germany). The gas flow rate was regulated by a mass flow controller (D07-19B, Beijing Sevenstar Electronics Co., Ltd., Beijing, China). Subsequently, the generated OAS were introduced into the venturi mixer to form the ROS solution via hydraulic cavitation. The ROS encompassed hydroxyl radical (•OH), superoxide radical (O2•−), singlet oxygen (1O2), and others [28]. The total concentration of ROS was quantified through the DPD (N, N-diethyl-p-phenylenediamine) colorimetric method, as specified in the USEPA Method 330.5. The relationship between ROS concentration and the gas–liquid ratio (σ) is illustrated in Figure S1.

2.2.2. Simulated Experiment of FLO Degradation

In order to investigate the kinetics and mechanisms of antibiotic degradation in this system, a laboratory-based simulated experiment was conducted using deionized water at a flow rate of 1.2 L/min. The concentration of FLO in the simulation was set to 1 mg/L, and reaction times ranging from 0.6 s to 20 s were examined. All other experimental parameters were consistent with those used in the aquaculture seawater treatment experiment.

2.3. Analytical Methods

2.3.1. Quantitation of Antibiotics

A liquid chromatography system (LC, Agilent 1260, Agilent Technologies, Santa Clara, CA, USA) coupled with mass spectrometry (MS, Agilent 6460, Agilent Technologies, Santa Clara, CA, USA) was used for the quantitative analysis of antibiotics. MS was carried out in the multi-reaction monitoring mode. The parameters for MS are detailed in Table 2.

2.3.2. Identification of Intermediates During FLO Degradation

To identify the intermediates generated during FLO degradation, 500 mL of collected samples were subjected to SPE for enrichment. Subsequently, the intermediates were analyzed by LC-MS. Additionally, a gas chromatography-mass spectrometry (GC-MS, Agilent 7890A, 5975C MSD, Agilent Technologies, Santa Clara, CA, USA) was used for the identification of intermediates following derivatization by N, O-bis-(trimethylsilyl) trifluoroacetamide (BSTFA).

2.3.3. Determination of Water Quality

All water quality parameters were detected according to the Chinese Sea Water Quality Standard (GB 3097-1997) [33]. ClO3− and BrO3− were measured with an ion chromatograph (Thermo 2100, ThermoFisher Scientific, Waltham, MA, USA). Trihalomethanes (THMs) were analyzed using a GC (Agilent 7890A, Agilent Technologies, Santa Clara, CA, USA) equipped with an electron capture detector. ClO3−, BrO3−, and THMs were detected according to the Chinese Sanitary Standards for Drinking Water Quality (GB 5749-2006) [34].

3. Results

3.1. Degradation of Antibiotics During Aquaculture Seawater Treatment

The aquaculture seawater was subjected to pretreatment via filtration, foam flotation, and microbiological processes before being mixed with eight antibiotics in the pipeline. Subsequently, a solution containing ROS was introduced into the pipeline to degrade the antibiotics, with a final σ of 0.012. Samples were collected in bottles containing sodium thiosulfate at a reaction time of 20 s. The variations in antibiotic concentration in aquaculture water before and after treatment are shown in Figure 2 and Figure S2.
The initial concentrations of OTC, TC, CTC, SDZ, SMZ, SMX, CHL, and FLO were 104.42 μg/L, 94.70 μg/L, 115.16 μg/L, 118.04 μg/L, 112.09 μg/L, 100.88 μg/L, 99.76 μg/L, and 99.40 μg/L, respectively. Following a 20-s treatment with the ROS solution, all eight antibiotics were reduced to levels below the detection limit, demonstrating that the OAHC-based system efficiently and rapidly degrades antibiotics in aquaculture water.

3.2. The Effect of OAHC-Based System on the Water Quality

The water quality data for the effluents from aquaculture tanks, biological filter, and post-OHC-based system treatment are listed in Table 3. The water temperature, salinity, and sulfide levels remained relatively stable throughout the treatment process. The pH value exhibited a slight increase following the OAHC-based system treatment, likely attributable to the oxidation of certain acidic organic compounds present in the water [35]. Dissolved oxygen concentrations rose from 9.79 mg/L to 13.19 mg/L post-treatment with the OAHC-based system. Given that the self-purification of water bodies necessitates the consumption of dissolved oxygen [36], an elevated dissolved oxygen level is more favorable for mitigating the adverse effects of discharged aquaculture effluent on surrounding aquatic environments.
The turbidity of the effluent from the aquaculture tanks was initially measured at 1.44 NTU. Following pretreatment, this value decreased to 0.62 NTU and was further reduced to 0.27 NTU after treatment with the OAHC-based system. Turbidity serves as an indicator of light obstruction in a liquid, primarily influenced by the scattering effect of suspended solids and the absorption effect of solutes [37]. After filtration and foam flotation processes, most of the suspended solids were effectively removed, suggesting that subsequent reductions in turbidity were mainly attributed to the degradation of organic molecules in the water. This indicates that the OAHC-based system can efficiently degrade recalcitrant organic matter, thereby enhancing water quality. Consistent with these findings, the chemical oxygen demand (CODMn), which characterizes the concentration of organic matter in water, also showed a significant reduction—from 2.71 mg/L in the initial effluent to 2.12 mg/L after pretreatment, and finally to 1.84 mg/L following OAHC-based system treatment.
Ammonia nitrogen and nitrite nitrogen are significant constituents of inorganic nitrogen. Excessive discharge into the environment can lead to seawater eutrophication. The concentrations of ammonia nitrogen and nitrite nitrogen in aquaculture effluent were 1.20 mg/L and 0.18 mg/L, respectively. After pretreatment, these concentrations decreased to 0.43 mg/L and 0.14 mg/L, respectively. It is evident that the reduction patterns of ammonia nitrogen and nitrite nitrogen differed throughout the process. Ammonia nitrogen was predominantly removed during pretreatment, while the OAHC-based system had a limited effect on it. Conversely, nitrite nitrogen remained relatively stable during pretreatment and was primarily removed through OAHC-based system treatment. During pretreatment, ammonia nitrogen was mainly reduced via nitrification, with nitrite nitrogen acting as an intermediate product that was continuously produced and removed, thereby maintaining a relatively stable concentration.
The primary chemical by-products generated during the ozone oxidation process are bromate (BrO3−) and trihalomethanes, both of which are recognized as potential carcinogens that can accumulate in organisms, thereby posing a potential risk to human health [38]. Neither bromate nor dichlorobromomethane was detected before or after treatment with the OAHC-based system. However, the concentrations of chlorodibromomethane and tribromomethane increased from 0.01 μg/L and 0.03 μg/L, respectively, to 0.13 μg/L and 4.87 μg/L post-treatment. This increase is attributed to the reactivity of ozone with Cl- and Br- ions in water, leading to the formation of secondary oxidants such as hypochlorite (OCl−) and hypobromite (OBr−), which possess oxidizing properties and react with organic matter in water to generate trihalomethanes. Given that seawater is generally weakly alkaline, under alkaline conditions, the oxidation potentials of OCl− and OBr− are 0.90 V and 0.76 V, respectively. Consequently, OCl− can react with Br- to form OBr−, leading to the predominant formation of brominated by-products. Notably, the concentrations of these by-products are consistently maintained well below the limits prescribed by the “Chinese Sanitary Standards for Drinking Water Quality” (GB 5749-2006), indicating a minimal impact of the OAHC-based system on water quality.

3.3. Kinetic of FLO Degradation

To simulate the process of antibiotic degradation in the OAHC-based system, FLO was selected as the target antibiotic for the kinetic and mechanism studies due to its superior stability during ozonation compared to other antibiotics used in this study [21,39]. The variation in FLO concentration during the OAHC-based process under different σ is shown in Figure 3a. Approximately 70% of FLO degradation occurred immediately following miscibility within 0.6 s, with the remaining portion continuing to degrade during transport through the pipeline. The removal kinetics of FLO through the pipeline can be described by pseudo-first-order kinetics, similar to many organic compounds [40,41,42], as shown in Equation (1):
ln(ct/c0) = kt,
where ct (mg/L) is the concentration of FLO at time t (s), c0 (mg/L) is the initial concentration, and k (min−1) is the observed degradation rate constant.
As shown in Figure 3b, when the initial point is excluded, the degradation rate constant increases from 0.07 to 0.17 min−1 for σ from 0.006 to 0.024. Based on these results, the degradation process of FLO should comprise two stages: (1) direct degradation by ROS generated during the miscible process under hydraulic cavitation; and (2) continuous oxidation facilitated by residual ozone in the water.

3.4. Mechanism of FLO Degradation

To investigate the mechanism of FLO degradation in the OAHC-based system, the predominant intermediates were identified by LC-MS and GC-MS, as illustrated in Figure 4 and Figure 5, respectively. Based on the LC-MS chromatogram, a prominent peak corresponding to FLO was observed at 30 min in the control sample. When σ was set within the range of 0.006 to 0.012, a total of five distinct intermediates were identified. At σ = 0.024, the chromatogram nearly overlapped with the baseline, and no significant peaks were detected. The structural formulas of these intermediates were confirmed through MS/MS analysis by comparing their mass spectral data with previously reported findings [43,44]. The mass spectra of intermediates I (molecular mass 273), II (353), III (355), IV (335), and V (367) are shown in Figures S3–S7. The chromatogram obtained via GC-MS exhibits a similar trend to that acquired through LC-MS, with an additional three peaks (VI–VIII) being identified. The structural formulas of intermediates VI–VIII were identified by comparing the results with standard spectra from the mass spectrometry database (NIST 11), as depicted in Figures S8–S10.
Based on the analysis of intermediates, the proposed pathway for FLO degradation in the OAHC-based system is illustrated in Figure 6. Generally, FLO degradation in the OAHC-based system can be attributed to three primary pathways: (a) hydroxyl radical (•OH) targets highly electronegative sites, including fluorine, chlorine, as well as sulfomethyl groups, thereby initiating substitution reactions; (b) cleavage of the C13-N16 bond, and (c) the electrophilic attacking of •OH at the site of the benzene ring, forming a phenolic product. The final intermediate product identified in this study was intermediate VIII, which undergoes oxidation to form a structure analogous to catechol. This compound is subsequently oxidized to yield 1,2-benzoquinone. In aqueous solution, 1,2-benzoquinone readily undergoes ring-opening reactions and is ultimately oxidized to CO2 [45].

4. Discussion

As a potent oxidant with an oxidation potential of 2.07 V, O3 is widely utilized for sterilization and disinfection in aquaculture water. Owing to its ability to resonate into two distinct molecular structures, O3 exhibits both electrophilic and nucleophilic characteristics [46]. Numerous studies have been conducted on the degradation of antibiotics by ozone in recent years. Won et al. [47] introduced ozone gas with a rate of 10 mg/min into 1 L of river water to degrade three kinds of sulfonamide antibiotics, including SMX, achieving a degradation rate of 75–85% after 30 min. Hossain et al. [48] introduced ozone gas (flow rate: 3.0 L/min; ozone concentration: 0.55 mg/L) through a porous glass into 220 mL of solution containing OTC, achieving approximately 99% or greater OTC degradation within 30 min. Choi et al. [21] investigated the degradation of antibiotics (2 μmol/L), including OTC and FLO, using varying concentrations of O3 over a reaction period of 1 min. Complete degradation of OTC was achieved at an O3 concentration of 41 μmol/L, whereas the degradation efficiency of FLO was less than 10%, even at elevated O3 concentrations up to 104 μmol/L. Compared to these conventional ozone methods, the OAHC-based system shows a higher efficiency, with a reaction time as brief as 20 s.
Even when compared to other ozone-based advanced oxidation technologies, the OAHC-based system still demonstrates superior performance (as shown in Table S1). Guo et al. [49] introduced ozone gas at a rate of 15.36 mg/min into an SDZ solution under UV irradiation (with light intensity stabilized at 0.3 mW/cm2), achieving complete degradation of 25 mg/L SDZ within 7 min. Luo et al. [50] achieved a degradation efficiency of 93.2% for 30 mg/L TC within 25 min by introducing ozone gas at a rate of 0.4 mg/min using a CoSO catalyst. Pelalak et al. [51] attained an 89.34% degradation efficiency for 10 mg/L SMX within 40 min by introducing ozone gas at a rate of 5 mg/h into a 0.1 mmol/L H2O2 solution. Wang et al. [52] utilized an ultrafine bubble-generating device to introduce ozone at a rate of 0.08g/h, producing microbubbles with diameters ranging from 0.5 μm to 3 μm, and achieved a degradation efficiency of 98.7% for 50 mg/L TC within 20 min.
The oxidation of organic contaminants in O3 systems encompasses both direct oxidation by ozone molecules and indirect oxidation via ROS, including hydroxyl radical (•OH), superoxide radical (O2•−), singlet oxygen (1O2), among others [28]. Notably, •OH can adequately oxidize most organic contaminants in a short period without selectivity due to its higher redox potential (2.73 V) compared to other reactive oxygen species, such as 2.2 V for 1O2 and 0.89 V for O2•− [53,54,55]. The high efficiency of the OAHC-based system can be attributed to its abundant generation of OAS based on the strong ionization discharge in combination with the micro-streamer and micro-glow discharges, ultimately resulting in the formation of ROS, particularly •OH, in water [56]. Additionally, the gas–liquid mass transfer efficiency of ozone into water also affects the final treatment effect [57]. The hydraulic cavitation induced by a venturi mixer can produce microbubbles, which in turn can facilitate the formation of free radicals upon the collapse of these bubbles [29].
The reaction rate constant for the interaction between •OH and organic compounds generally falls within the range of 108 to 1010 L/mol·s [58]. In comparison, the reaction rate constant between •OH and ammonia nitrogen is 9.7 × 107 L/mol·s [59], which is 1–2 orders of magnitude lower than that for the reaction with organic matter. Consequently, in the presence of organic matter, •OH is less likely to react with ammonia nitrogen. Additionally, some organic compounds containing amino groups may release additional ammonia nitrogen during degradation, resulting in minimal changes in ammonia nitrogen concentration during •OH treatment in this study. Conversely, the reaction rate constant of nitrite nitrogen with •OH is 6.0 × 109 L/mol·s [60], facilitating rapid reactions between nitrite nitrogen and •OH.
BrO3− formation during oxidative treatment of bromide-containing drinking water has been of great concern ever since bromate was classified as potentially carcinogenic by the IARC (International Agency for the Research on Cancer) in 1990 [61]. Reducing ozonation duration may serve as an effective strategy for controlling the formation of BrO3− during the ozonation process, while it has less impact on the formation potentials of brominated trihalomethanes [62]. Therefore, the high efficiency of the OAHC-based system plays a crucial role in suppressing the formation of BrO3−.
Energy efficiency assessment is an essential component that plays a pivotal role in evaluating the feasibility of promoting and implementing a new system. The standalone O3 system has a power rating of 350 W and is capable of generating an ozone concentration of 110 mg/L at a flow rate of 1 L/min. With a σ of 0.012, the system can achieve a water treatment capacity of 5 m3/h, resulting in an energy efficiency of 0.07 kW·h/m3 for the O3 system. Indeed, the pump constitutes the primary energy consumer within this system, as a pump operating at a flow rate of 5 m3/h typically requires a power output in the range of 5 to 10 kW. This estimate is preliminary. When scaling up the implementation, numerous complex factors must be considered. For reference, when the OAHC-based system was employed for drinking water treatment with a capacity of 500 m3/h [31], an energy efficiency of 0.119 kW·h/m3 was measured. Among ozone-based advanced oxidation processes, the OAHC-based system exhibits economic competitiveness [63].
Overall, the OAHC-based system demonstrates significant potential for aquaculture seawater treatment. However, it is currently constrained by certain limitations, including its inability to effectively remove ammonia and the formation of brominated trihalomethanes. In the future, integrating OAHC-based systems with biotreatment technologies such as constructed wetlands and microbial treatment processes may emerge as promising application directions, which is expected to compensate for the limitations inherent in individual methods.

5. Conclusions

In recent years, O3-AOPs have attracted extensive attention due to the strong oxidation and relative environmental friendliness of ozone. However, the high costs and substantial energy requirements associated with these processes restrict their large-scale industrial applications. An OAHC-based system was developed for the efficient treatment of aquaculture seawater contaminated with antibiotics, which shows high efficiency. The main conclusions are as follows:
(1)
Based on the combination of strong ionization discharge and hydraulic cavitation, the OAHC-based system exhibits excellent degradation efficiency for antibiotics within a reaction time of 20 s.
(2)
Following treatment by the OAHC-based system, all water quality parameters met the limits specified in the Chinese Sea Water Quality Standard, and the DBPs complied with the China National Standards for drinking water quality.
(3)
The degradation process of FLO can be divided into two stages: the initial direct degradation by ROS generated during the miscible process under hydraulic cavitation, followed by continuous oxidation promoted by residual ozone in the water.
(4)
The FLO degradation in the OAHC-based system can be attributed to three primary pathways: (a) substitution reaction at highly electronegative sites; (b) cleavage of the C-N bond; and (c) the electrophilic attacking at the site of the benzene ring.

Supplementary Materials

The following supporting information can be downloaded at: https://www.mdpi.com/article/10.3390/w17040566/s1, Figure S1: The relationship between the ROS concentration and the gas-liquid ratio; Figure S2: The concentration of antibiotics in aquaculture water before and after treatment; Figure S3: Spectra of Intermediate I with LC-MS; Figure S4: Spectra of Intermediate II with LC-MS; Figure S5: Spectra of Intermediate III with LC-MS; Figure S6: Spectra of Intermediate IV with LC-MS; Figure S7: Spectra of Intermediate V with LC-MS; Figure S8: Spectra and structure of Intermediate VI with GC-MS; Figure S9: Spectra and structure of Intermediate VII with GC-MS; Figure S10: Spectra and structure of Intermediate VIII with GC-MS; Table S1: Comparison of antibiotic degradation by advanced oxidized technologies.

Author Contributions

Conceptualization, X.H. and D.Y.; methodology, X.H.; validation, D.Y., L.S., and Y.J.; formal analysis, X.H.; investigation, X.H.; resources, X.H.; data curation, X.H.; writing—original draft preparation, X.H.; writing—review and editing, D.Y., L.S., and Y.J.; visualization, X.H.; supervision, D.Y., L.S., and Y.J.; project administration, X.H.; funding acquisition, D.Y. All authors have read and agreed to the published version of the manuscript.

Funding

This research was funded by the Science and Technology Project of Huadong Engineering (Fujian) Corporation, grant number KY2023-FJ-02-01.

Data Availability Statement

Data are contained within the article and Supplementary Materials.

Acknowledgments

The authors gratefully thank Mindong Bai for providing the ozone generator based on the strong ionization discharge in combination with micro-streamer and micro-glow discharges, as well as for her guidance in exploring the •OH generation mechanism.

Conflicts of Interest

Authors Xiaodian Huang, Dong Yang, Liang Song and Yongcan Jiang were employed by the company Power China Huadong Engineering Corporation Ltd. All the authors declare that the research was conducted in the absence of any commercial or financial relationships that could be construed as a potential conflict of interest.

Abbreviations

The following abbreviations are used in this manuscript:
ARBsAntibiotic-resistant bacteria
AOPsAdvanced oxidation processes
CHLChloramphenicol
CTCChlorotetracycline
FLOFlorfenicol
GCGas chromatography
LCLiquid chromatography
MSMass spectrometry
OAHCOzone assisted with hydrodynamic cavitation
OASOxygen activated species
OTCOxytetracycline
ROSReactive oxygen species
SDZSulfadiazine
SMXSulfamethoxazole
SMZSulfamerazine
SPESolid-phase extraction
TCTetracycline

References

  1. Campanati, C.; Willer, D.; Schubert, J.; Aldridge, D.C. Sustainable Intensification of Aquaculture through Nutrient Recycling and Circular Economies: More Fish, Less Waste, Blue Growth. Rev. Fish. Sci. Aquac. 2022, 30, 143–169. [Google Scholar] [CrossRef]
  2. The State of World Fisheries and Aquaculture 2024; FAO: Rome, Italy, 2024; ISBN 978-92-5-138763-4.
  3. Henriksson, P.J.G.; Belton, B.; Jahan, K.M.-; Rico, A. Measuring the Potential for Sustainable Intensification of Aquaculture in Bangladesh Using Life Cycle Assessment. Proc. Natl. Acad. Sci. USA 2018, 115, 2958–2963. [Google Scholar] [CrossRef]
  4. Schar, D.; Klein, E.Y.; Laxminarayan, R.; Gilbert, M.; Van Boeckel, T.P. Global Trends in Antimicrobial Use in Aquaculture. Sci. Rep. 2020, 10, 21878. [Google Scholar] [CrossRef] [PubMed]
  5. Yan, F.; An, L.; Xu, X.; Du, W.; Dai, R. A Review of Antibiotics in Surface Water and Their Removal by Advanced Electrocoagulation Technologies. Sci. Total Environ. 2024, 906, 167737. [Google Scholar] [CrossRef] [PubMed]
  6. Hirsch, R.; Ternes, T.; Haberer, K.; Kratz, K.L. Occurrence of Antibiotics in the Aquatic Environment. Sci. Total Environ. 1999, 225, 109–118. [Google Scholar] [CrossRef]
  7. McCorquodale-Bauer, K.; Grosshans, R.; Zvomuya, F.; Cicek, N. Critical Review of Phytoremediation for the Removal of Antibiotics and Antibiotic Resistance Genes in Wastewater. Sci. Total Environ. 2023, 870, 161876. [Google Scholar] [CrossRef]
  8. Chen, L.; Li, H.; Liu, Y.; Cui, Y.; Li, Y.; Yang, Z. Distribution, Residue Level, Sources, and Phase Partition of Antibiotics in Surface Sediments from the Inland River: A Case Study of the Xiangjiang River, South-Central China. Environ. Sci. Pollut. Res. Int. 2020, 27, 2273–2286. [Google Scholar] [CrossRef]
  9. Feng, L.; Cheng, Y.; Zhang, Y.; Li, Z.; Yu, Y.; Feng, L.; Zhang, S.; Xu, L. Distribution and Human Health Risk Assessment of Antibiotic Residues in Large-Scale Drinking Water Sources in Chongqing Area of the Yangtze River. Environ. Res. 2020, 185, 109386. [Google Scholar] [CrossRef] [PubMed]
  10. Chen, Y. Dynamic Evolution of Antibiotic Resistance Genes in Plastisphere in the Vertical Profile of Urban Rivers. Water Res. 2024, 249, 120946. [Google Scholar] [CrossRef] [PubMed]
  11. Yang, Y.; Song, W.; Lin, H.; Wang, W.; Du, L.; Xing, W. Antibiotics and Antibiotic Resistance Genes in Global Lakes: A Review and Meta-Analysis. Environ. Int. 2018, 116, 60–73. [Google Scholar] [CrossRef] [PubMed]
  12. Zainab, S.M.; Junaid, M.; Xu, N.; Malik, R.N. Antibiotics and Antibiotic Resistant Genes (ARGs) in Groundwater: A Global Review on Dissemination, Sources, Interactions, Environmental and Human Health Risks. Water Res. 2020, 187, 116455. [Google Scholar] [CrossRef]
  13. Liu, K. Occurrence and Source Identification of Antibiotics and Antibiotic Resistance Genes in Groundwater Surrounding Urban Hospitals. J. Hazard. Mater. 2024, 465, 133368. [Google Scholar] [CrossRef] [PubMed]
  14. Zheng, D.; Yin, G.; Liu, M.; Chen, C.; Jiang, Y.; Hou, L.; Zheng, Y. A Systematic Review of Antibiotics and Antibiotic Resistance Genes in Estuarine and Coastal Environments. Sci. Total Environ. 2021, 777, 146009. [Google Scholar] [CrossRef]
  15. Shen, J. Land Use Conversion to Uplands Significantly Increased the Risk of Antibiotic Resistance Genes in Estuary Area. Environ. Int. 2024, 191, 108953. [Google Scholar] [CrossRef] [PubMed]
  16. Xu, J.-Y.; Ding, J.; Du, S.; Zhu, D. Tire Particles and Its Leachates: Impact on Antibiotic Resistance Genes in Coastal Sediments. J. Hazard. Mater. 2024, 465, 133333. [Google Scholar] [CrossRef] [PubMed]
  17. Wang, H.; Wang, N.; Wang, B.; Zhao, Q.; Fang, H.; Fu, C.; Tang, C.; Jiang, F.; Zhou, Y.; Chen, Y.; et al. Antibiotics in Drinking Water in Shanghai and Their Contribution to Antibiotic Exposure of School Children. Environ. Sci. Technol. 2016, 50, 2692–2699. [Google Scholar] [CrossRef] [PubMed]
  18. Li, N.; Ho, K.W.K.; Ying, G.G.; Deng, W.J. Veterinary Antibiotics in Food, Drinking Water, and the Urine of Preschool Children in Hong Kong. Environ. Int. 2017, 108, 246–252. [Google Scholar] [CrossRef] [PubMed]
  19. He, L.-X.; He, L.-Y.; Gao, F.-Z.; Zhang, M.; Chen, J.; Jia, W.-L.; Ye, P.; Jia, Y.-W.; Hong, B.; Liu, S.-S.; et al. Mariculture Affects Antibiotic Resistome and Microbiome in the Coastal Environment. J. Hazard. Mater. 2023, 452, 131208. [Google Scholar] [CrossRef] [PubMed]
  20. Von Gunten, U. Oxidation Processes in Water Treatment: Are We on Track? Environ. Sci. Technol. 2018, 52, 5062–5075. [Google Scholar] [CrossRef]
  21. Choi, S.; Sim, W.; Jang, D.; Yoon, Y.; Ryu, J.; Oh, J.; Woo, J.-S.; Kim, Y.M.; Lee, Y. Antibiotics in Coastal Aquaculture Waters: Occurrence and Elimination Efficiency in Oxidative Water Treatment Processes. J. Hazard. Mater. 2020, 396, 122585. [Google Scholar] [CrossRef] [PubMed]
  22. Huang, A.; Yan, M.; Lin, J.; Xu, L.; Gong, H.; Gong, H. A Review of Processes for Removing Antibiotics from Breeding Wastewater. Int. J. Environ. Res. Public Health 2021, 18, 4909. [Google Scholar] [CrossRef] [PubMed]
  23. Yang, W.; Li, J.; Yao, Z.; Li, M. A Review on the Alternatives to Antibiotics and the Treatment of Antibiotic Pollution: Current Development and Future Prospects. Sci. Total Environ. 2024, 926, 171757. [Google Scholar] [CrossRef] [PubMed]
  24. Chávez, A.M.; Gimeno, O.; Rey, A.; Pliego, G.; Oropesa, A.L.; Álvarez, P.M.; Beltrán, F.J. Treatment of Highly Polluted Industrial Wastewater by Means of Sequential Aerobic Biological Oxidation-Ozone Based AOPs. Chem. Eng. J. 2019, 361, 89–98. [Google Scholar] [CrossRef]
  25. Liu, H.; Gao, Y.; Wang, J.; Ma, D.; Wang, Y.; Gao, B.; Yue, Q.; Xu, X. The Application of UV/O3 Process on Ciprofloxacin Wastewater Containing High Salinity: Performance and Its Degradation Mechanism. Chemosphere 2021, 276, 130220. [Google Scholar] [CrossRef] [PubMed]
  26. Chen, H.; Wang, J. Degradation and Mineralization of Ofloxacin by Ozonation and Peroxone (O3/H2O2) Process. Chemosphere 2021, 269, 128775. [Google Scholar] [CrossRef] [PubMed]
  27. Rosenfeldt, E.J.; Linden, K.G.; Canonica, S.; Von Gunten, U. Comparison of the Efficiency of OH Radical Formation during Ozonation and the Advanced Oxidation Processes O3/H2O2 and UV/H2O2. Water Res. 2006, 40, 3695–3704. [Google Scholar] [CrossRef] [PubMed]
  28. Dai, M.; Niu, Q.; Wu, S.; Lin, Y.; Biswas, J.K.; Yang, C. Hydroxyl Radicals in Ozone-Based Advanced Oxidation of Organic Contaminants: A Review. Environ. Chem. Lett. 2024, 22, 3059–3106. [Google Scholar] [CrossRef]
  29. Sakr, M.; Mohamed, M.M.; Maraqa, M.A.; Hamouda, M.A.; Aly Hassan, A.; Ali, J.; Jung, J. A Critical Review of the Recent Developments in Micro–Nano Bubbles Applications for Domestic and Industrial Wastewater Treatment. Alex. Eng. J. 2022, 61, 6591–6612. [Google Scholar] [CrossRef]
  30. Bai, M.; Zheng, W.; Huang, X.; Wang, Y.; Li, H.; Zhang, Z. Studies on the ·OH Inactivation of Non-Native Aquatic Organisms and Potential Disinfection Byproduct Formation for Oceanic Environmental Safety. Plasma Chem. Plasma Process. 2018, 38, 1051–1062. [Google Scholar] [CrossRef]
  31. Bai, M.; Huang, X.; Zhong, Z.; Cao, M.; Gao, M. Comparison of OH and NaClO on Geosmin Degradation in the Process of Algae Colonies Inactivation at a Drinking Water Treatment Plant. Chem. Eng. J. 2020, 393, 123243. [Google Scholar] [CrossRef]
  32. Bai, M.; Bai, X.; Zhang, Z.; Bai, M.; Yang, B. Treatment of Red Tide in Ocean Using Non-Thermal Plasma Based Advanced Oxidation Technology*. Plasma Chem. Plasma Process. 2005, 25, 539–550. [Google Scholar] [CrossRef]
  33. GB 3097-1997; Sea Water Quality Standard. Ministry of Ecology and Environment of the People’s Republic of China: Beijing, China, 1997.
  34. GB 5749-2006; Sanitary Standards for Drinking Water Quality. National Health Commission of the People’s Republic of China: Beijing, China, 2006.
  35. Ochir, D.; Lee, Y.; Shin, J.; Kim, S.; Kwak, J.; Chon, K. Oxidative Treatments of Pesticides in Rainwater Runoff by HOCl, O3, and O3/H2O2: Effects of pH, Humic Acids and Inorganic Matters. Separations 2021, 8, 101. [Google Scholar] [CrossRef]
  36. Akatah, B.; Izinyon, C. Self-Purification Capacity of Mmubete Stream in Rivers State, Nigeria. IOSR J. Mech. Civil. Eng. 2023, 20, 37–45. [Google Scholar] [CrossRef]
  37. Elevli, S.; Uzgören, N.; Bingöl, D.; Elevli, B. Drinking Water Quality Control: Control Charts for Turbidity and pH. J. Water Sanit. Hyg. Dev. 2016, 6, 511–518. [Google Scholar] [CrossRef]
  38. Simpson, A.M.-A.; Mitch, W.A. Chlorine and Ozone Disinfection and Disinfection Byproducts in Postharvest Food Processing Facilities: A Review. Crit. Rev. Environ. Sci. Technol. 2022, 52, 1825–1867. [Google Scholar] [CrossRef]
  39. Ben, W.; Qiang, Z.; Pan, X.; Nie, Y. Degradation of Veterinary Antibiotics by Ozone in Swine Wastewater Pretreated with Sequencing Batch Reactor. J. Environ. Eng. 2012, 138, 272–277. [Google Scholar] [CrossRef]
  40. Hermes, N.; Jewell, K.S.; Falås, P.; Lutze, H.V.; Wick, A.; Ternes, T.A. Ozonation of Sitagliptin: Removal Kinetics and Elucidation of Oxidative Transformation Products. Environ. Sci. Technol. 2020, 54, 10588–10598. [Google Scholar] [CrossRef] [PubMed]
  41. Tang, L.; Zhou, S.; Li, F.; Sun, L.; Lu, H. Ozone Micronano-Bubble-Enhanced Selective Degradation of Oxytetracycline from Production Wastewater: The Overlooked Singlet Oxygen Oxidation. Environ. Sci. Technol. 2022, 57, acs.est.2c06008. [Google Scholar] [CrossRef]
  42. Li, B. Effects of Common Inorganic Anions on the Ozonation of Polychlorinated Diphenyl Sulfides on Silica Gel: Kinetics, Mechanisms, and Theoretical Calculations. Water Res. 2020, 186, 116358. [Google Scholar] [CrossRef] [PubMed]
  43. Zhang, Y.; Li, J.; Zhou, L.; Wang, G.; Feng, Y.; Wang, Z.; Yang, X. Aqueous Photodegradation of Antibiotic Florfenicol: Kinetics and Degradation Pathway Studies. Environ. Sci. Pollut. Res. Int. 2016, 23, 6982–6989. [Google Scholar] [CrossRef] [PubMed]
  44. Liu, N.; Sijak, S.; Zheng, M.; Tang, L.; Xu, G.; Wu, M. Aquatic Photolysis of Florfenicol and Thiamphenicol under Direct UV Irradiation, UV/H2O2 and UV/Fe(II) Processes. Chem. Eng. J. 2015, 260, 826–834. [Google Scholar] [CrossRef]
  45. Wang, J.; Xu, M.; Liang, X.; Zhang, Y.; Yang, D.; Pan, L.; Fang, W.; Zhu, C.; Wang, F. Development of a Novel 2D Ni-MOF Derived NiO@C Nanosheet Arrays Modified Ti/TiO2NTs/PbO2 Electrode for Efficient Electrochemical Degradation of Salicylic Acid Wastewater. Sep. Purif. Technol. 2021, 263, 118368. [Google Scholar] [CrossRef]
  46. Wei, C.; Zhang, F.; Hu, Y.; Feng, C.; Wu, H. Ozonation in Water Treatment the Generation, Basic Properties of Ozone and Its Practical Application. Rev. Chem. Eng. 2017, 33, 49–89. [Google Scholar] [CrossRef]
  47. Won, J.S.; Kaewsuk, J.; Jo, J.H.; Lim, D.-H.; Seo, G.T. A Density Functional Theory Study on the Ozone Oxidation of Sulfonamide Antibiotics. J. Adavanced Oxid. Technol. 2015, 18, 31–38. [Google Scholar] [CrossRef]
  48. Hossain, A.; Nishimura, Y.; Salma, U.; Tokumura, M.; Nishino, T.; Raknuzzaman, M.; Noro, K.; Watanabe, K.; Amagai, T.; Makino, M. Effects of Water Matrices on the Removal of Oxytetracycline Antibiotic and Total Organic Carbon (TOC) Using Four Different Oxidation Processes. Results Eng. 2024, 24, 103183. [Google Scholar] [CrossRef]
  49. Guo, W.; Yang, Z.; Du, J.; Yin, R.; Zhou, X.; Jin, S.; Ren, N. Degradation of Sulfadiazine in Water by a UV/O3process: Performance and Degradation Pathway. RSC Adv. 2016, 6, 57138–57143. [Google Scholar] [CrossRef]
  50. Luo, L.; Zou, D.; Lu, D.; Yu, F.; Xin, B.; Ma, J. Study of Catalytic Ozonation for Tetracycline Hydrochloride Degradation in Water by Silicate Ore Supported Co3O4. RSC Adv. 2018, 8, 41109–41116. [Google Scholar] [CrossRef] [PubMed]
  51. Pelalak, R.; Alizadeh, R.; Ghareshabani, E.; Heidari, Z. Degradation of Sulfonamide Antibiotics Using Ozone-Based Advanced Oxidation Process: Experimental, Modeling, Transformation Mechanism and DFT Study. Sci. Total Environ. 2020, 734, 139446. [Google Scholar] [CrossRef] [PubMed]
  52. Wang, C.; Lin, C.-Y.; Liao, G.-Y. Degradation of Antibiotic Tetracycline by Ultrafine-Bubble Ozonation Process. J. Water Process Eng. 2020, 37, 101463. [Google Scholar] [CrossRef]
  53. Nazaroff, W.W.; Cass, G.R. Mathematical Modeling of Chemically Reactive Pollutants in Indoor Air. Environ. Sci. Technol. 1986, 20, 924–934. [Google Scholar] [CrossRef] [PubMed]
  54. Wardman, P. Reduction Potentials of One-Electron Couples Involving Free Radicals in Aqueous Solution. J. Phys. Chem. Ref. Data 1989, 18, 1637–1755. [Google Scholar] [CrossRef]
  55. Wang, Y.; Lin, Y.; He, S.; Wu, S.; Yang, C. Singlet Oxygen: Properties, Generation, Detection, and Environmental Applications. J. Hazard. Mater. 2024, 461, 132538. [Google Scholar] [CrossRef] [PubMed]
  56. Bai, M.; Zhang, Z.; Xue, X.; Yang, X.; Hua, L.; Fan, D. Killing Effects of Hydroxyl Radical on Algae and Bacteria in Ship’s Ballast Water and on Their Cell Morphology. Plasma Chem. Plasma Process. 2010, 30, 831–840. [Google Scholar] [CrossRef]
  57. Huang, X.; Quan, X.; Cheng, W.; Cheng, C.; Cheng, Z.; Yang, L.; Jiang, L. Enhancement of Ozone Mass Transfer by Stainless Steel Wire Mesh and Its Effect on Hydroxyl Radical Generation. Ozone-Sci. Eng. 2020, 42, 347–356. [Google Scholar] [CrossRef]
  58. Wols, B.A.; Hofman-Caris, C.H.M. Review of Photochemical Reaction Constants of Organic Micropollutants Required for UV Advanced Oxidation Processes in Water. Water Res. 2012, 46, 2815–2827. [Google Scholar] [CrossRef] [PubMed]
  59. Hickel, B.; Sehested, K. Reaction Of Hydroxyl Radicals with Ammonia In Liquid Water at Elevated-Temperatures. Radiat. Phys. Chem. 1992, 39, 355–357. [Google Scholar] [CrossRef]
  60. Ocampo-Pérez, R.; Rivera-Utrilla, J.; Mota, A.J.; Sánchez-Polo, M.; Leyva-Ramos, R. Effect of Radical Peroxide Promoters on the Photodegradation of Cytarabine Antineoplastic in Water. Chem. Eng. J. 2016, 284, 995–1002. [Google Scholar] [CrossRef]
  61. Von Gunten, U.; Oliveras, Y. Advanced Oxidation of Bromide-Containing Waters: Bromate Formation Mechanisms. Environ. Sci. Technol. 1998, 32, 63–70. [Google Scholar] [CrossRef]
  62. Lin, T.; Wu, S.; Chen, W. Formation Potentials of Bromate and Brominated Disinfection By-Products in Bromide-Containing Water by Ozonation. Environ. Sci. Pollut. Res. Int. 2014, 21, 13987–14003. [Google Scholar] [CrossRef] [PubMed]
  63. Katsoyiannis, I.A.; Canonica, S.; Von Gunten, U. Efficiency and Energy Requirements for the Transformation of Organic Micropollutants by Ozone, O3/H2O2 and UV/H2O2. Water Res. 2011, 45, 3811–3822. [Google Scholar] [CrossRef] [PubMed]
Water 17 00566 g001
Figure 1. Sketch of the system for aquaculture seawater treatment.
Figure 1. Sketch of the system for aquaculture seawater treatment.
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Water 17 00566 g002
Figure 2. Chromatogram of antibiotics in aquaculture water before and after treatment (Abbreviations: ND, not detected).
Figure 2. Chromatogram of antibiotics in aquaculture water before and after treatment (Abbreviations: ND, not detected).
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Water 17 00566 g003
Figure 3. The variation in FLO concentration during the OAHC-based process (a) and pseudo-first-order kinetics (b) under different σ.
Figure 3. The variation in FLO concentration during the OAHC-based process (a) and pseudo-first-order kinetics (b) under different σ.
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Figure 4. Chromatogram of FLO degradation in the OAHC-based system with different gas–liquid ratios by LC-MS.
Figure 4. Chromatogram of FLO degradation in the OAHC-based system with different gas–liquid ratios by LC-MS.
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Water 17 00566 g005
Figure 5. Chromatogram of FLO degradation in the OAHC-based system with different gas–liquid ratios by GC-MS.
Figure 5. Chromatogram of FLO degradation in the OAHC-based system with different gas–liquid ratios by GC-MS.
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Figure 6. Pathways of FLO degradation in the OAHC-based system.
Figure 6. Pathways of FLO degradation in the OAHC-based system.
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Table 1. Details of six antibiotics used in this study.
Table 1. Details of six antibiotics used in this study.
ClassAntibioticsMolecular StructureMolecular Mass (g/mol)
TetracyclinesTetracycline
(TC)		Water 17 00566 i001444.43
Oxytetracycline
(OTC)		Water 17 00566 i002460.43
Chlorotetracycline
(CTC)		Water 17 00566 i003478.88
SulfonamidesSulfadiazine
(SDZ)		Water 17 00566 i004250.28
Sulfamerazine
(SMZ)		Water 17 00566 i005264.30
Sulfamethoxazole
(SMX)		Water 17 00566 i006253.23
PhenicolsChloramphenicol
(CHL)		Water 17 00566 i007323.13
Florfenicol
(FLO)		Water 17 00566 i008358.21
Table 2. Parameter for six antibiotics detected by mass spectrometer.
Table 2. Parameter for six antibiotics detected by mass spectrometer.
AntibioticsPrecursor IonTransitions (m/z)Fragmentor (V)Collision Energy (V)
TC[M + H]+445→41012020
OTC[M + H]+461→42612020
CTC[M + H]+479→44412020
SDZ[M + H]+251→9210030
SMZ[M + H]+265→9212030
SMX[M + H]+254→9211030
CHL[M − H]+321→2571305
FLO[M − H]+356→3361405
Table 3. Water quality parameters of aquaculture wastewater before and after treatment.
Table 3. Water quality parameters of aquaculture wastewater before and after treatment.
Detected ItemsAquaculture Tank EffluentBiological Filter EffluentEffluent After OAHCLimit Values
Temperature (°C)24.224.323.9—
Salinity (‰)34.835.134.9—
pH7.207.357.867.8~8.5 1
Dissolved oxygen (mg/L)9.309.7913.19>5 1
Turbidity (NTU)1.440.620.27—
CODMn (mg/L)2.712.121.84≤3 1
Ammonia nitrogen (mg/L)1.200.430.44—
Nitrite nitrogen (mg/L)0.180.140.01—
Sulfide (mg/L)0.040.030.04≤0.05 1
Bromate (μg/L)ND—ND≤10 2
Trichloromethane (μg/L)0.12—0.16≤60 2
Bromodichloromethane (μg/L)0.01—0.03≤60 2
Dibromochloromethane (μg/L)0.01—0.13≤100 2
Tribromomethane (μg/L)0.03—4.87≤100 2
Notes: Abbreviations: ND, not detected; —: no data. 1 GB 3097-1997, Chinese Sea Water Quality Standard (second category); 2 GB 5749-2006, Chinese Sanitary Standards for Drinking Water Quality.
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Huang, X.; Yang, D.; Song, L.; Jiang, Y. Degradation of Antibiotics in Aquaculture Seawater: A Treatment Based on Ozone Assisted with Hydrodynamic Cavitation. Water 2025, 17, 566. https://doi.org/10.3390/w17040566

AMA Style

Huang X, Yang D, Song L, Jiang Y. Degradation of Antibiotics in Aquaculture Seawater: A Treatment Based on Ozone Assisted with Hydrodynamic Cavitation. Water. 2025; 17(4):566. https://doi.org/10.3390/w17040566

Chicago/Turabian Style

Huang, Xiaodian, Dong Yang, Liang Song, and Yongcan Jiang. 2025. "Degradation of Antibiotics in Aquaculture Seawater: A Treatment Based on Ozone Assisted with Hydrodynamic Cavitation" Water 17, no. 4: 566. https://doi.org/10.3390/w17040566

APA Style

Huang, X., Yang, D., Song, L., & Jiang, Y. (2025). Degradation of Antibiotics in Aquaculture Seawater: A Treatment Based on Ozone Assisted with Hydrodynamic Cavitation. Water, 17(4), 566. https://doi.org/10.3390/w17040566

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