An Experimental Study on the Novel Ozone-Electro-Fenton Coupled Reactor for Treating Ofloxacin-Containing Industrial Wastewater

Abstract

Industrial organic wastewater, with its complex composition, high biological toxicity, and recalcitrance, has become a major challenge in water pollution control. This is especially true for antibiotic-containing wastewater, such as ofloxacin wastewater, for which there is an urgent need to develop effective treatment technologies. Conventional treatment processes are insufficiently efficient, while individual advanced oxidation processes (AOPs) have drawbacks such as poor oxidation selectivity and catalyst deactivation. To address these issues, researchers have explored the coupling of different AOPs and found that such combinations can enhance the oxidation performance, achieve complementary advantages, reduce the equipment costs, and offer great development potential. An experiment was conducted to evaluate the performance of an Ozone-Electro-Fenton coupled process in treating ofloxacin industrial wastewater. The results demonstrated that under the same conditions, after four hours of treatment, the coupled process achieved a 70% reduction in the UV absorption peak of the wastewater, compared to less than 20% for individual processes, indicating a significant synergistic effect. Further optimization of the ozone aeration structure revealed that with a hole size of 0.5 mm, single-layer aeration holes, and six holes, the COD removal rate reached 96% after six hours, the ozone utilization improved to 85%, and the gas holdup stabilized at 4.6%. Under these conditions, the mixture of ozone and air bubbles formed mixed bubbles. Influenced by the electric field and electrode plate wall effects, the bubble residence time was prolonged. The bubble size was approximately 2.8 mm, the gas flow horizontal velocity was about 18.5 m/s, and after a horizontal displacement of 0.17 mm in the wastewater, the lateral velocity became zero. The ratio of the distance between the bubble center and the wall to the equivalent bubble diameter was approximately 3.45. The bubbles were subject to a strong wall effect, which extended their residence time. This not only facilitated the removal of small bubbles from the electrode plates but also enhanced the ion diffusion near the plates, thereby boosting pollutant degradation. This study shows that the Ozone-Electro-Fenton coupled process is highly effective in degrading ofloxacin industrial wastewater, offering an innovative solution for treating other antibiotic-containing wastewater. Future research will focus on further optimizing the process, improving its adaptability to complex matrix wastewater, and validating it at the pilot scale to promote its engineering application.

1. Introduction

Many types of industrial wastewater exist, with pharmaceutical wastewater being a typical representative due to its prevalence and recalcitrance. Its treatment is of utmost importance. Ofloxacin (OFL), a quinolone antibiotic with broad-spectrum antibacterial activity, is widely used in human medicine and animal husbandry to combat various bacterial infections. During the COVID-19 pandemic, its usage in clinical settings has significantly increased [1]. Ofloxacin has the chemical formula C₁₈H₂₀FN₃O₄, a molecular weight of 361.36 g/mol, and a CAS number of 83934-44-0. It is usually a white to pale yellow crystalline powder with a melting point of 188–192 °C, slightly water-soluble but miscible with organic solvents like ethanol and acetone, and highly chemically stable, making it persistent in the environment.

In both human and veterinary medicine, ofloxacin holds a significant position. It effectively treats a range of bacterial infections, including respiratory and urinary tract infections such as pneumonia, cystitis, and pyelonephritis. In veterinary clinical practice, it is commonly used to treat bacterial diseases in livestock and aquatic animals. However, as humans and animals absorb less than 10% of the drug, most of it is excreted in urine and feces, leading to environmental pollution. This issue has been further highlighted during the COVID-19 pandemic due to the substantial increase in ofloxacin usage, which has increased its release into the environment.

Ofloxacin has emerged as a significant pollutant, being detected in various environmental media such as wastewater, surface water, groundwater, drinking water, and sludge. Its high chemical stability makes it resistant to natural degradation in water and soil, allowing it to persist and pose ongoing ecological risks. It can also accumulate in organisms and biomagnify through the food chain, indirectly affecting human health.

Moreover, ofloxacin exhibits acute and chronic toxicity to aquatic life, affecting fish growth, reproduction, and survival, and disrupting the balance of aquatic ecosystems. It can also inhibit the growth and reproduction of soil biota such as earthworms and nematodes, alter soil biodiversity, and impair soil ecological functions and fertility. Additionally, ofloxacin residues can induce antibiotic resistance genes in microorganisms. These genes can spread among different microbes via horizontal gene transfer, enabling bacteria to develop a resistance to ofloxacin. This increases the difficulty of treating infections and poses a serious threat to public health. Prolonged or excessive exposure to ofloxacin may lead to health issues such as allergies and gut microbiota imbalance in humans, and can also cause damage to vital organs like the liver and kidneys [2,3,4]. Its specific properties are summarized in the following

Table 1. Attributes of ofloxacin.

Attribute	Details
Chemical Formula	C18H20FN3O4
Molecular Weight	361.36 g/mol
Physical State	White to Pale Yellow Crystalline Powder
Melting Point	188–192 °C
Environmental Stability	Chemically stable and resistant to degradation in water and soil
Bioaccumulation	Easily absorbed and accumulated by organisms, with potential for food chain transfer

:

The treatment methods for ofloxacin wastewater mainly include conventional treatment technologies and advanced oxidation processes. Conventional treatment technologies have significant drawbacks in treating antibiotic-containing wastewater. For example, adsorption in physical methods is unstable, and microbial activity in biological methods is inhibited [5,6]. To enhance conventional treatments, researchers have explored the modification of adsorption materials, the development of new composite materials, and the identification of antibiotic-resistant bacteria [7,8,9,10,11,12,13,14,15]. However, conventional technologies are less effective than advanced oxidation technologies (AOPs) in treating antibiotic-containing wastewater [16,17].

AOPs degrade pollutants efficiently by generating highly reactive hydroxyl radicals (·OH) in situ and offer significant advantages over traditional physicochemical methods. Their core mechanism involves activating oxidants using energy inputs (thermal, optical, or electrical) to rapidly mineralize organic matter into CO₂, H₂O, and inorganic salts under mild conditions (room temperature and pressure). AOPs include Fenton oxidation, ozonation, photocatalytic oxidation (UV/H₂O₂), electrochemical oxidation, and ultrasonic oxidation. They feature fast degradation rates, high mineralization degrees, and compact reactor designs [18,19].

Fenton oxidation, a key AOP, uses Fe²⁺ and H₂O₂ in acidic conditions to generate ·OH radicals for organic pollutant mineralization. Its effectiveness hinges on maintaining the Fe²⁺/Fe³⁺ redox cycle. However, it requires strict pH control (2.5–3.5) and faces secondary pollution from iron sludge [20,21,22,23].

The Electro-Fenton process combines traditional Fenton oxidation with electrochemistry, enabling efficient in-situ generation of reactive species like ·OH and catalyst regeneration. This improves the oxidation efficiency, reduces the chemical consumption, and decreases the iron sludge production. It has been successfully applied in the decolorization of dyestuff wastewater and the detoxification of pharmaceutical wastewater.

For instance, Zare et al. used a Ti/RuO₂ anode and Fe-Fe₂O₂ catalyst in Electro-Fenton experiments on ciprofloxacin (CIP). At pH 8.83, with a 14.80-min reaction time, 19.19 mA/cm² current density, 15.13 mg/L initial CIP concentration, and 199.03 mg/L catalyst dosage, they achieved 100% CIP removal and 45% total organic carbon (TOC) removal [24].

Despite its advantages, the Electro-Fenton process faces challenges in large-scale applications due to a low current efficiency, high electrode material costs, and unresolved mass transfer issues in reactor design.

Researchers have also explored combined and coupled processes. Ren et al. developed a cathodic membrane filtration (CMF) electrochemical reactor with a Ti/SnO₂-Sb anode and titanium mesh cathode, integrating membrane and electrochemical technologies for phosphorus removal. At 4 A/m² current density, 16.6 L/m² h membrane flux, and a Ca/P molar ratio of 1.67, they achieved a 96.2% phosphorus removal with an energy consumption of 45.7 kWh/kg P, showing an improvement over standalone electrochemical treatment [25].

Batool et al. reported a study on using manganese dioxide (MnO₂) as an earth-abundant, eco-friendly electrocatalyst to achieve over 99% mineralization of triclosan (TCS) and other halogenated phenols at parts-per-million (ppm) levels in wastewater. The researchers’ fabrication of two highly active MnO₂ phases—α-MnO₂-CC and δ-MnO₂-CC—on carbon cloth (CC) supported and evaluated their performance in oxidatively degrading TCS under pH-neutral conditions, including in simulated chlorinated wastewater, real wastewater, and both synthetic and real landfill leachates. A total organic carbon (TOC) analysis confirmed an effective TCS degradation, while electron paramagnetic resonance (EPR) and ultraviolet–visible (UV-Vis) spectroscopy identified reactive oxygen species (ROS), enabling the construction of a detailed TCS degradation pathway. Upon optimization, the TCS removal rate reached 38.38 nmol min⁻¹, outperforming previously reported rates using precious or toxic metal co-catalysts [26].

Lagum et al. coupled membrane bioreactors with anaerobic ammonium oxidation to overcome the limitations of separate membrane bioreactors in treatment performance and low-temperature tolerance. The coupled process reduced the footprint and costs and provided a reference for using the technology in cold regions [27].

Ozone-Electro-Fenton (OEF) combines ozonation and Electro-Fenton processes. Ozone reacts with H₂O₂ generated at the Electro-Fenton cathode to produce additional ·OH and HO₂· radicals, compensating for the insufficient radical generation in Electro-Fenton alone. Ozone also reacts with Fe²⁺ to form highly reactive ·O₃⁻ intermediates, which further convert to ·OH, creating a radical cascade amplification effect [28]. Compared to the single ozonation or Electro-Fenton process, OEF operates effectively over a wider pH range and is suitable for various water qualities. It efficiently degrades refractory organics and is ideal for treating high-toxicity, complex wastewater from pharmaceutical, chemical, and dye industries. In practice, OEF reduces chemical usage and resource optimization, lowering treatment costs while improving efficiency and stability. Future research can explore its adaptability and optimization in different industrial wastewater treatments for broader engineering applications.

Research on bubble motion characteristics in industrial wastewater treatment has made significant progress, particularly regarding the movement of near-wall bubbles. Studies show that the wall distance and the initial bubble shape are key parameters controlling the bubble motion [29]. The closer the bubble to the wall, the more pronounced the interference with its upward velocity and trajectory. Also, the initial bubble shape significantly alters its motion behavior. Additionally, the jet effect induced by near-wall bubbles and their group interactions are crucial driving factors for flow field dynamics [30]. The jet effect enhances local flow turbulence, promoting gas-liquid mass transfer, altering the bubble distribution and trajectories, and impacting the overall flow field structure.

It has been pointed out that the application of an external electric field has a significant regulatory effect on the bubble behavior [31,32]. The electric field changes the charge distribution on the bubble surface, inducing bubble deformation. This effectively increases the specific surface area and enhances the gas-liquid mass transfer efficiency. Such deformation not only strengthens the bubble’s mass transfer ability but also alters its trajectory, thereby optimizing the local flow field’s turbulence and improving the mass transfer efficiency.

The research on ozone bubbles mainly focuses on the application of microbubble technology. Taking drinking water treatment as an example, ozone microbubble aeration technology can not only significantly enhance the removal of organic matter but also effectively suppress the formation of bromate [33]. In the treatment of complex systems such as coking wastewater and high-salt dye wastewater, ozone microbubble technology also demonstrates excellent pollutant removal performance [34]. However, there remains a gap in the research on the movement of ozone bubbles in the near-wall region and their coupling effects under the action of an electric field.

In this study, a novel Ozone-Electro-Fenton coupled reactor was used to treat ofloxacin wastewater. Firstly, it generates more types and a larger amount of free radicals. Secondly, the electrode plates in the electro-Fenton process act as baffles. This design effectively prolongs the ozone residence time and improves its utilization efficiency. In a different way to previous studies mainly focusing on ozone microbubble technology, this study originally explores the motion of ozone bubbles near the wall and their coupling interaction with electric fields. By clarifying the internal mechanisms of the relevant processes, this research aims to provide new theoretical support for optimizing the Ozone-Electro-Fenton system and enhancing degradation efficiency. This innovative approach is expected to fill existing technological gaps, promote the development of industrial wastewater treatment technologies, and lay a foundation for more efficient and stable wastewater treatment solutions.

2. Materials and Methods

2.1. Experimental Materials

2.1.1. Experimental Water

The specific water quality parameters of the ofloxacin wastewater are detailed in

Table 2. Quality of ofloxacin wastewater.

Parameter	Value
pH	7.0 ± 0.2
COD (mg/L)	5120 ± 100
Ofloxacin Concentration (mg/L)	150 ± 10

.

2.1.2. Experimental Reagents

The reagents used in the experiment are listed in

Table 3. List of experimental drugs.

Name	Purity	Manufacturer	Location
Potassium Dichromate	AR	Tianjin Damo Chemical Reagent Factory	Tianjin, China
Sulfuric Acid	AR	Tianjin Damo Chemical Reagent Factory	Tianjin, China
Potassium Chromate	AR	Beijing Bligh & Weber Technology Co., Ltd.	Beijing, China
Mercuric Sulfate	AR	Shanghai Merida Biochemical Technology Co., Ltd.	Shanghai, China
Potassium Iodide	AR	Beijing Bligh & Weber Technology Co., Ltd.	Beijing, China
Methyl Orange	AR	Beijing Bligh & Weber Technology Co., Ltd.	Beijing, China
Malachite Green	AR	Beijing Bligh & Weber Technology Co., Ltd.	Beijing, China
Sodium Hydroxide	AR	Beijing Bligh & Weber Technology Co., Ltd.	Beijing, China
Potassium Titanium Oxalate	AR	Shanghai Dibai Biotechnology Co., Ltd.	Shanghai, China
Catalase	-	Beijing Bligh & Weber Technology Co., Ltd.	Beijing, China

.

2.1.3. Experimental Instruments

The instruments used in the experiment are listed in

Table 4. Experimental instruments.

Instrument Name	Model	Manufacturer	Location
UV Spectrophotometer	INESAL5UV	Shanghai Yidian Scientific Instruments Co., Ltd.	Shanghai, China
Gas Rotameter	LBZ-10WB	Shunda Lai Measurement and Control Equipment Co., Ltd.	Nanjing, China
Electric Hotplate Stirrer	79-1	Changzhou Jimatai Instruments Factory	Changzhou, China
Ozone Generator	FG-L	Guangzhou Feige Environmental Protection Technology Co., Ltd.	Guangzhou, China
pH Meter	PHS-3C	Shanghai Yuake Instruments Co., Ltd.	Shanghai, China
Electronic Balance	FA2004	ShanghaiYidian Scientific Instruments Co., Ltd.	Shanghai, China
DC Power Supply	MS-305D	Shanghai Xuxin Electrical Technology Co., Ltd.	Shanghai, China

.

2.2. Experimental Setup and Methods

2.2.1. Experimental Setup

The experimental setup, as shown in Figure 1, consists of an ozone supply system, an electrochemical reactor, and a tail gas treatment unit. The ozone system includes an ozone generator (with a maximum ozone output of 6 g/h and adjustable power) and a rotameter, connected to a gas distributor via a PTFE tube. The reactor, made of organic glass (internal dimensions 240 mm × 120 mm × 120 mm), houses alternately arranged graphite anodes (100 mm × 100 mm × 1 mm) and corrugated iron cathodes (with a single fold length of 20 mm, an angle of 80°, and overall dimensions of 140 mm × 100 mm × 1 mm). The aeration system comprises a PVC pipe network (internal diameter 4 mm) linked to the ozone supply system for gas distribution control. A DC power supply (0–5 A adjustable current range) drives the electrochemical reaction. Unreacted ozone and tail gases are treated via a serial KI absorption device to ensure environmental safety. The arrangement of the electrode plates and aeration pipes is shown in Figure 2.

2.2.2. Experimental Methods

1.

Electro-Fenton Oxidation Method

During the experiment, first accurately measure 1500 mL of wastewater and inject it into the reactor. Then, start the ozone generator without activating the ozone production function and set the air flow rate. Simultaneously, turn on the DC power supply and adjust it to the desired current value. After completing these steps, quickly start the timing device. Throughout the experiment, take samples at predetermined intervals. After collecting the samples, filter them and further treat with catalase before subsequent analysis. Adjust the pH value using dilute sulfuric acid and sodium hydroxide solution.

2.

Ozone Oxidation Method

For this method, first accurately measure 1500 mL of wastewater and inject it into the reactor. Turn on the ozone generator and activate the ozone production function. Carefully adjust valves 1 and 2 while closely monitoring the flowmeter readings to finely regulate the ozone flow until it stabilizes. Once the ozone flow is stable, open valve 3 to allow the gas to be evenly distributed through the distributor and flow into the aeration pipes. At this point, start the timing device. During the experiment, take samples at preset intervals. Filter the collected samples and perform subsequent analyses. To ensure experimental safety and effectively treat residual gases, use a potassium iodide (KI) solution to absorb the residual ozone throughout the experiment. The electrode plate spacing is set at 20 mm based on the piping requirements and relevant literature, and the pH is maintained at neutral [35,36].

3.

Ozone-Electro-Fenton Coupled Oxidation Method

For this procedure, first accurately measure 1500 mL of wastewater and inject it into the reactor. Then, start the ozone generator and activate the ozone production function. Adjust the ozone generation rate by controlling the power knob on the ozone generator and set the appropriate gas flow rate. At the same time, turn on the DC power supply and adjust it to the target current value. After completing these steps, immediately start the timing device. During the experiment, take samples at preset intervals. Filter the collected samples and further treat them with catalase before analysis. Adjust the pH value using a dilute sulfuric acid and sodium hydroxide solution. To ensure experimental safety and effectively treat residual gases, use a potassium iodide (KI) solution to absorb the residual ozone throughout the experiment.

2.3. Analytical Methods and Data Processing

1.

Chemical Oxygen Demand (COD)

In this experiment, the COD values are measured using a UV spectrophotometer to determine the absorbance of Cr³⁺ generated from the reduction of Cr⁶⁺ in a potassium dichromate solution at a specific wavelength. The COD values are directly proportional to the increase in the absorbance, and the sample COD values are calculated accordingly [37]. As a key indicator of the wastewater quality, the COD removal efficiency is assessed using Equation (1) to evaluate the degradation efficiency of the ofloxacin wastewater by the treatment process.

$$\eta ={\displaystyle \frac{c_0-c_t}{c_0}}\times 100\%$$

(1)

where c₀ and c_t represent the initial COD and COD at time t, respectively.

2.

pH Value

A pH meter is utilized to measure the pH of the experimental water samples.

3.

Absorbance

Ultraviolet spectrophotometry is employed to analyze the test substances. Specific molecules in the samples absorb radiation in the 200–700 nm spectral range, and a UV spectrophotometer measures the absorbance at a specific wavelength. This method is highly sensitive and simple to operate, making it suitable for detecting various substances.

4.

Ozone Concentration Measurement

The ozone concentration is determined using the iodometric method. Ozone is introduced into a solution containing potassium iodide and sulfuric acid, oxidizing potassium iodide to iodine. Subsequently, a sodium thiosulfate standard solution is used to titrate to the endpoint, and the ozone concentration is calculated based on the titration consumption. To prevent interference from H₂O₂, separate detection is required.

5.

Gas Holdup Measurement

The volume expansion method is used to measure the gas holdup in the liquid [38]. Based on the principle that the liquid volume changes as gas dissolves and is released, the calculation Formula (2) is as follows:

$${\epsilon }_g=\left(1-{\displaystyle \frac{V_0}{V}}\right)\times 100\%=\left(1-{\displaystyle \frac{H_0}{H}}\right)\times 100\%$$

(2)

where ε_g is the gas holdup, V₀ is the volume of water before the gas is introduced, V is the volume of water after the gas is introduced, H₀ is the effective water level before the gas is introduced, and $\mathrm{H}$ is the effective water level after the gas is introduced.

6.

H₂O₂ Detection

Ozone and iron ions can interfere with common iodimetric and permanganate methods. Therefore, the potassium titirconal acetate spectrophotometric method is employed for qualitative analysis of H₂O₂ in this experiment [39]. Under acidic conditions, H₂O₂ reacts with the potassium titirconal acetate to form a stable orange complex, which is used to detect the presence of H₂O₂ in the solution.

3. Results

Experimental studies on treating ofloxacin industrial wastewater were conducted using a novel Ozone-Electro-Fenton coupled reactor. The focus was on treatment effectiveness and the ozone aeration structure.

3.1. Ozone-Electro-Fenton Coupled Treatment of Ofloxacin Industrial Wastewater: Effectiveness

Experiments were conducted to treat ofloxacin industrial wastewater using ozone, Electro-Fenton, and Ozone-Electro-Fenton coupled processes. Figure 3 shows the degradation results of the ofloxacin. After 4 h, the Ozone-Electro-Fenton process achieved a UV absorption peak decline of over 70%, surpassing the results of the other processes.

3.2. Influence of Ozone Aeration Structure

In the Ozone-Electro-Fenton coupled treatment process, in addition to the impact of the process parameters on the treatment effectiveness, the aeration structure of the ozone could affect the movement and residence time of the ozone bubbles in water, which in turn significantly influenced the oxidation and degradation of organic pollutants. Therefore, this study focuses on the ozone aeration structure, particularly examining the effects of parameters such as the hole size, number of holes, and hole spacing [40].

Response Surface Methodology Analysis

Response surface methodology was employed to investigate the effects of the hole size (X1), the number of holes (X2), and the hole spacing (X3) on the UV absorption peak decline rate, as well as the relationships among these three factors, aiming to identify the appropriate aeration structure. A schematic diagram of the hole structure is shown in Figure 4, where the holes completely penetrate the tube. The independent variables were set as the hole size (X1), the number of holes (X2), and the hole spacing (X3), with the response value being the UV absorption peak decline rate in ofloxacin wastewater after 40 min of treatment. A total of 17 runs and 5 replicate experiments were conducted to eliminate pure error.

Table 5. BBD levels and independent values.

Factor	Independent Variable	Level
		−1	0	+1
X1	Hole size (mm)	0.5	1.5	2.5
X2	Number of holes	2	3	4
X3	Hole spacing (mm)	4	6	8

presents the levels and independent values of the Box-Behnken design (BBD).

Experiments were conducted based on the process conditions provided by the Design-Expert 13. The regression equation derived from the second-order polynomial is shown in Equation (3).

$$\begin{array}{ll}y=& 4.49624-75.63211X_1+66.27799\times X_2+8.96822\times X_3\\ & \hspace{1em}\hspace{1em}\hspace{1em}\hspace{1em}+ 3.15675\times X_1\times X_2-2.52804\times X_1\times X_3\\ & \hspace{1em}\hspace{1em}\hspace{1em}\hspace{1em}- 0.462112\times X_2\times X_3+17.809851\times X_1^2-12.44250\times X_2^2\\ & \hspace{1em}\hspace{1em}\hspace{1em}\hspace{1em}- 0.642831\times X_3^2\end{array}$$

(3)

The experimental results and variance analysis were presented in

Table 6. Experimental and predicted UV absorption peak decline rate under different conditions.

Run	Independent Variables			Response
	X1	X2	X3	UV Absorption Peak Decline Rate
				Experimental Value (%)	Predicted Value (%)
1	0.5	2	6	74.7482	74.6133
2	2.5	2	6	10.3177	12.4987
3	0.5	4	6	57.6517	55.4707
4	2.5	4	6	5.8482	5.9831
5	0.5	3	4	77.743	77.7054
6	2.5	3	4	34.37	93.6908
7	0.5	3	8	69.7674	72.1209
8	2.5	3	8	6.1701	6.20771
9	1.5	2	4	29.9264	30.0989
10	1.5	4	4	16.8996	19.11182
11	1.5	2	8	18.4694	16.2508
12	1.5	4	8	1.7457	1.57319
13	1.5	3	6	31.894	31.7741
14	1.5	3	6	28.601	31.7741
15	1.5	3	6	35.4906	31.7741
16	1.5	3	6	29.3888	31.7741
17	1.5	3	6	33.4961	31.7741

and

Table 7. ANOVA results and adequacy.

Source	Sum of Squares	Df	Mean Square	F-Value	p-Value
Model	9111.30	9	1012.37	112.33	<0.0001
X1	6227.52	1	6227.52	691.01	<0.0001
X2	329.17	1	329.17	36.53	0.0005
X3	492.77	1	492.77	54.68	0.0002
X1X2	39.86	1	39.86	4.42	0.0735
X1X3	102.26	1	102.26	11.35	0.0119
X2X3	3.42	1	3.42	0.3791	0.5576
X12	1335.54	1	1335.54	148.19	<0.0001
X22	651.86	1	651.86	72.33	<0.0001
X32	27.84	1	27.84	3.09	0.1222
Residual	63.09	7	0.0493
LOF	30.54	3	0.0610	1.51	0.4026
Pure Error	32.55	4	0.0405
SD = 3.00, Mean = 33.09, R2 = 0.9931, Adj.R2 = 0.9843, Pred. R2 = 0.9412, AP = 33.0656

.

In this model, the F-value is 112.33 and the p-value is less than 0.0001, indicating the model’s reliability. The R² of 0.9931 shows it can explain 99.31% of data variability, suggesting a great fit. Even after adjusting for the number of independent variables, the Adjusted R² remains high at 0.9843. The Predicted R² of 0.9412 also indicates good predictive power. The experimental and predicted values show a basically linear relationship (Figure 5). Furthermore, the non-significant Lack of Fit test result demonstrates a good fit between the model and experimental data. Among hole size, number of holes, and hole spacing, their significance order is hole size > hole spacing > number of holes.

3.3. Interaction Analysis

As shown in Figure 6, the effects of the hole size and number of holes on the UV absorption peak decline rate of ofloxacin wastewater after 40 min of treatment were investigated with the hole spacing fixed at 6 mm. The curved shape of the surface in the figure indicated that the relationship was not a simple linear one but rather a nonlinear interaction. This suggests a complex interplay between the hole size and the number of holes. The contour analysis revealed that the hole size had a more significant impact on the UV absorption peak decline rate of ofloxacin wastewater compared to the number of holes, which was consistent with the results of the analysis and calculations. When the hole size is reduced from 2.5 mm to 0.5 mm and the number of holes is increased from 2 to 3, the UV absorption peak decline rate increases. However, if the number of holes is further increased, the UV absorption peak decline rate starts to decline. This might be due to the fact that as the number of holes increases, the total area of the aeration holes also increases, which can help distribute the gas more evenly. But when there are too many holes, a larger proportion of the gas is released from higher positions, leading to a reduced residence time in the water and thus a decrease in the treatment effectiveness.

As shown in Figure 7, the effects of the hole size and hole spacing on the UV absorption peak decline rate of ofloxacin wastewater after 40 min of treatment were studied with the number of holes fixed at 3. The curved shape of the surface in the figure indicates that the relationship is not a simple linear one but rather a nonlinear interaction. This suggests a complex interplay between the hole size and hole spacing. The contour analysis revealed that the hole size had a more significant impact on the UV absorption peak decline rate of ofloxacin wastewater compared to the hole spacing, which was consistent with the results of the analysis and calculations. When the hole size is increased from 0.5 mm to 2.5 mm and the hole spacing is increased from 4 mm to 8 mm, the UV absorption peak decline rate decreases. The reason for the influence of the hole spacing on the UV absorption peak decline rate may be that as the hole spacing increases, the height of the aeration holes above increases, the gas is released at a higher position, the residence time in the water is reduced, and thus the treatment effectiveness decreases.

As shown in Figure 8, the effects of the number of holes and hole spacing on the UV absorption peak decline rate of the ofloxacin wastewater after 40 min of treatment were investigated with the hole size fixed at 1.5 mm. The curved shape of the surface in the figure indicates that the relationship is not a simple linear one but rather a nonlinear interaction. This suggests a complex interplay between the number of holes and hole spacing. The contour analysis reveals that the hole spacing has a more significant impact on the UV absorption peak decline rate of the ofloxacin wastewater compared to the number of holes, which is consistent with the results of the analysis and calculations. When the number of holes is increased from 2 to 3, the UV absorption peak decline rate increases. The UV absorption peak decline rate remains high when the hole spacing is between 4 mm and 5.5 mm. Although the contour lines do not form a complete ellipse, the optimal hole spacing is around 5 mm based on the overall data. Subsequently, single-factor experiments on hole spacing and the UV absorption peak decline rate of the ofloxacin wastewater were conducted to further verify the reliability, and the results were basically consistent.

From the analysis of the experimental data in Figure 6, Figure 7 and Figure 8, we chose a configuration with a hole size of 0.5 mm, a hole spacing of 5 mm, and three through-holes (i.e., six holes). Under these conditions, the UV absorption peak decline rate of the ofloxacin wastewater reached 79% after 4 h of treatment. However, during the experiment, we observed that as the number of holes increased, the position of some holes became too high, causing the discharged bubbles to have a shorter residence time. To avoid the impact of different heights, we conducted experiments with 0.5 mm holes at the same height, as shown in Figure 9.

Subsequent experimental results showed that when three through-holes (i.e., six holes) with a diameter of 0.5 mm were opened at the same height on a tube, the UV absorption peak decline rate of ofloxacin wastewater reached 82% after four hours of treatment, surpassing the configurations with holes at different heights. This indicates that opening holes at the same height can prolong the residence time of the bubbles and thus enhance wastewater treatment efficiency.

Based on the above-mentioned optimized conditions, tests were conducted on the gas holdup, ozone utilization efficiency, and COD removal efficiency. The results demonstrated that compared to the ozone process, the Ozone-Electro-Fenton coupled process significantly increased the gas holdup from 2% to 4.6% and the ozone utilization efficiency from 34% to 85%. Furthermore, continuous monitoring of the COD degradation over six hours revealed a substantial improvement in the COD removal efficiency. Figure 10 illustrates the detailed results.

As shown in the figure, the COD removal rate gradually stabilized after the fifth hour of the experiment, reaching 95.7% by the sixth hour. Notably, similar to the trend of the UV absorption peak decline rate, the COD removal rate was not significant within the first hour of the experiment. This may be because, although ofloxacin is decomposed in a short time, the primary products of its decomposition still have a high COD value, which need to further breakdown to significantly reduce the COD level.

3.4. Flow Characteristics Analysis

The gas-liquid flow in the novel Ozone-Electro-Fenton reactor is complex. Near the electrode plates, electrolysis generates tiny bubble clusters. These bubbles, under the electric field’s influence, rise slowly and may even adhere to the plates, hindering anode reactions and oxygen diffusion to the cathode. The ozone aeration tube between the plates releases ozone bubbles, which are much larger than the tiny bubbles near the plates and form a continuous flow of individual large bubbles. The main flow characteristic is the interaction between these large ozone bubbles and the tiny bubble clusters near the plates. This interaction can be categorized into two scenarios:

Moderate Horizontal Velocity of Ozone Bubbles: When the horizontal velocity of ozone bubbles from the aeration tube is moderate, they form a zigzag upward flow near the electrode plates. This interaction increases turbulence, dislodges some tiny bubbles, and enhances oxygen and ferrous ion diffusion. The anode promotes the detachment of tiny bubble clusters, accelerating bubble rise and escape, while the cathode disrupts the gas film on the electrode plates, boosting the generation of hydrogen peroxide and hydroxyl radicals, thereby enhancing coagulation and oxidation processes and improving ozone utilization.

High Horizontal Velocity of Ozone Bubbles: When the ozone bubbles have a high horizontal velocity, they encounter significant resistance. Upon exiting the aeration tube, they rise and escape immediately, failing to reach the electrode plate region. As a result, they cannot disturb the microbubble clusters adhering to the plates. The ozone bubbles rise rapidly, leading to a low gas holdup of ozone in the wastewater, which is detrimental to the Ozone-Electro-Fenton process.

Figure 11 illustrates these two bubble rise scenarios.

From the experimental results, the flow of the ozone bubbles in the novel Ozone-Electro-Fenton reactor includes all the mentioned scenarios. In practice, some bubbles collide with the electrode plates, which is beneficial for mass transfer. The horizontal velocity of ozone bubbles is primarily determined by the bubble diameter and the ozone gas flow rate. The ozone bubble diameter is primarily determined by the orifice size of the ozone aeration tube. If the ozone gas flow rate is constant, the horizontal velocity of the ozone bubbles is mainly determined by the total orifice area.

The diameter of ozone bubbles is the main factor affecting the bubble flow. The orifice size of the ozone aeration tube determines the size of the bubbles. For ozone bubbles, it can be considered as a continuous jet flow of individual bubbles. To simplify calculations, the equivalent diameter of a single bubble is used, as shown in Equation (4).

$${\mathrm{d}}_{\mathrm{e}\mathrm{q}}=d_b\approx 2·{\left({\displaystyle \frac{3d_{hole}\sigma }{4\left({\rho }_l-{\rho }_g\right)g}}\right)}^{{\displaystyle \frac{1}{3}}}$$

(4)

In the formula: d_eq is the equivalent diameter of a bubble, in meters (m).

d_b is the detachment diameter of a single bubble, in meters (m).

d_hole is hole size, in meters (m).

σ is the liquid surface tension. For wastewater, it’s approximately $0.072 \mathrm{N}/\mathrm{m}$.

ρ_l is the liquid phase density. For wastewater, it’s roughly $1000 \mathrm{k}\mathrm{g}/{\mathrm{m}}^3$.

ρ_g is the gas phase density, which is about $1.2 \mathrm{k}\mathrm{g}/{\mathrm{m}}^3$.

$\mathrm{g}$ is the gravitational acceleration, taken as $9.81 \mathrm{m}/{\mathrm{s}}^2$.

When the orifice diameters are 0.5 mm, 1.5 mm, and 2.5 mm, the corresponding equivalent bubble diameters calculated from the above formula are approximately 2.8 mm, 4 mm, and 4.7 mm, respectively, all within the range of 0.7–25 mm. According to literature studies, bubbles in this diameter range have a zigzag rising trajectory [41,42].

If the horizontal velocity of ozone bubbles is zero, whether the bubble rise is affected by the wall effect can be determined by Formulas (5)–(10) [43,44]. Calculations show that when the horizontal velocity of bubbles is zero, the residence time of gas in the reactor and the horizontal displacement of bubbles vary with different orifice diameters.

$$v_{gas}={\displaystyle \frac{Q}{A}}$$

(5)

$$Re={\displaystyle \frac{{\rho }_l{v}_{gas}{\mathrm{d}}_{\mathrm{e}\mathrm{q}}}{\mu }}$$

(6)

$$F_D={\displaystyle \frac{1}{2}}{C}_D{\rho }_l{v}_{gas}{A}_1$$

(7)

$${\mathrm{m}}_{\mathrm{e}\mathrm{q}}{\displaystyle \frac{d{v}_x}{dt}}=-F_D$$

(8)

$${\mathrm{m}}_{\mathrm{e}\mathrm{q}}={\displaystyle \frac{\pi {\mathrm{d}}_{\mathrm{e}\mathrm{q}}^3}{6}}{\rho }_g$$

(9)

$$t_x={\displaystyle \frac{2{\mathrm{m}}_{\mathrm{e}\mathrm{q}}}{C_D{\rho }_l{v}_{gas}{A}_1}}\ln \left({\displaystyle \frac{v_{gas}}{v_x}}\right)\left(take{ v}_x as {10}^{-6}\mathrm{m}/\mathrm{s}\right)$$

(10)

In the formula: ${\mathrm{v}}_{\mathrm{g}\mathrm{a}\mathrm{s}}$ is the horizontal velocity of bubbles, m/s.

Q is the total gas flow rate, $1.308\times {10}^{-4} {\mathrm{m}}^3$/s.

A is the total orifice area. With six holes per tube and 36 tubes in total, ${\mathrm{m}}^2$.

$\mathrm{R}\mathrm{e}$ is the Reynolds number.

μ is the liquid dynamic viscosity, taken as $0.001 \mathrm{P}\mathrm{a}·\mathrm{s}$.

F_D is the drag force, N.

${\mathrm{m}}_{\mathrm{e}\mathrm{q}}$ is the mass of a single bubble, kg.

t_x is the horizontal deceleration time, seconds (s).

v_x is the horizontal velocity of a bubble, m/s.

The terminal rise velocity of bubbles for each orifice diameter was calculated assuming a constant bubble diameter, as shown in Equation (11). This velocity can be used as the average rise speed of bubbles [45].

$$V_T=\sqrt{{\displaystyle \frac{4}{3C_D}}{\displaystyle \frac{{\rho }_l-{\rho }_g}{{\rho }_l}}g{d}_{eq}}$$

(11)

In the formula: V_T is the terminal rise velocity of a bubble, m/s.

The horizontal distance from the bubble center to the wall when the bubble’s horizontal velocity is zero is defined as L.

$$L^{*}=L/d_{eq}$$

(12)

When L* is less than 3.5, the bubbles are subject to a strong wall effect, causing their rise velocity to decrease by approximately 10% [46]. With L known to be about 4 mm and the liquid level height approximately 60 mm, and combined with relevant parameters under each orifice diameter, the results are shown in

Table 8. Parameters for each orifice diameter.

Hole Size mm	Gas Velocity m/s	Horizontal Displacement at Zero Horizontal Velocity mm	L* 1	Terminal Rise Velocity m/s	Total Residence Time s	Specific Surface Area m2/m3
0.5	18.5	0.17	3.45	0.259	0.231	2143
1.5	2.06	0.21	2.40	0.310	0.194	1500
2.5	0.74	0.23	2.03	0.336	0.178	1277

¹ L* is the ratio of the distance from the bubble center to the wall to the bubble diameter when the transverse velocity of the bubble is zero.

.

Based on the calculation results in Table 8, within the millimeter-scale orifice diameter range, the behavior of the ozone bubbles exhibits significant regularity: as the orifice diameter increases, the equivalent bubble diameter grows, leading to an increase in the terminal rise velocity of bubbles and a consequent shortening of residence time. Although the wall effect gradually intensifies with larger orifice diameters, the bubble residence time still becomes shorter.

The specific surface area of bubbles, a critical factor influencing mass transfer, is primarily affected by the bubble size. As shown in Table 8, the specific surface area decreases with an increasing bubble diameter, reducing the interfacial area for mass transfer per unit volume and thus decreasing gas-liquid mass transfer efficiency—an unfavorable outcome for mass transfer. Overall, an orifice diameter of 0.5 mm offers a more suitable comprehensive performance in terms of both the bubble specific surface area and residence time.

Increasing the number of orifices to evenly distribute ozone bubbles is an intuitive approach. Experiments with multi-layer orifices along the height of the ozone aeration tube showed that while more orifices enhanced gas-liquid mixing, ozone bubbles entering the electrode plate region from different heights shortened their residence time in the target area [47].

Additionally, orifice spacing affects mass transfer efficiency. Too small a spacing promotes bubble coalescence, reducing the specific surface area; too large a spacing significantly shortens the residence time of bubbles generated at higher positions, decreasing ozone utilization. When the vertical spacing between the orifices on the ozone aeration tube exceeds three times the bubble diameter, the bubble swarm exhibits a synergistic zigzag swinging trajectory during rise, with wake interference between adjacent bubbles effectively suppressed. This reduces gas-liquid turbulence, decreases gas-liquid mass transfer efficiency, and slightly shortens the total residence time of ozone bubbles—all detrimental to the Ozone-Electro-Fenton process [48]. Therefore, multi-layer orifice arrangements on the ozone aeration tube are generally unfavorable for mass transfer.

To address these issues, a single-layer bottom multi-orifice design for the ozone aeration tube is beneficial for uniform bubble distribution. This avoids positional differences along the tube’s height and effectively extends the total residence time of the bubbles in the target area.

Through the analysis of gas-liquid flow characteristics in the novel Ozone-Electro-Fenton reactor, the optimized ozone aeration structure promotes the diffusion of oxygen and ferrous ions generated at the anode via bubble motion. Oxygen diffusing to the cathode provides sufficient oxygen for hydrogen peroxide production, while ferrous ions diffusing to the cathode enhance flocculation, promote redox reactions, and remove some oxidizable pollutants. Additionally, ferrous ion diffusion homogenizes the ion concentration distribution on the cathode surface, reducing concentration polarization, improving current efficiency, and lowering energy consumption and operational costs [49].

4. Discussion and Conclusions

This study employed a novel Ozone-Electro-Fenton coupled reactor to systematically investigate the treatment performance of the Ozone-Electro-Fenton process for organic wastewater. Through process comparisons and an optimization of the ozone aeration structure, the synergistic enhancement between the ozone oxidation and electro-Fenton reactions was revealed. The main conclusions are as follows:

(1)

For actual ofloxacin industrial wastewater, experimental results showed that the Ozone-Electro-Fenton coupled process exhibited superior treatment efficiency compared to individual processes (ozone alone or electro-Fenton alone). Additionally, the influence of structural parameters of the ozone aeration system— the orifice diameter, hole spacing, and number of holes—on the ofloxacin degradation rate followed the significance order: orifice diameter > hole spacing > number of holes. The optimal ozone aeration structure (0.5 mm orifice diameter, single-layer configuration, six holes) achieved an 85% ozone utilization efficiency after 6 h, a cumulative COD removal rate of 96%, and a stable gas holdup of 4.6%.

(2)

In the novel Ozone-Electro-Fenton coupled reactor, under the dual effects of the electric field and wall effect, the ozone bubbles form wall-colliding and zigzag flows near the electrode plate surface. These bubbles carry small bubbles adhering to the wall to detach, disrupting the formation of the gas film on the electrode plate surface. This increases turbulence, promotes the diffusion of oxygen and ferrous ions to the cathode, enhances the generation of reactive species (e.g., hydroxyl radicals) at the cathode, and strengthens the oxidation and mass transfer processes. Simultaneously, it further prolongs bubble residence time and improves mass transfer efficiency.