Investigation on Mechanism of Tetracycline Removal from Wastewater by Sinusoidal Alternating Electro-Fenton Technique

Abstract

Sinusoidal alternating electro-Fenton (SAEF) is a new type of advanced electrochemical oxidation technology for the treatment of refractory organic wastewater. In this research, the removal performance and degradation mechanism of tetracycline (TC) were investigated, and the optimal operation parameters were determined. Scanning electron microscope (SEM), X-ray diffraction (XRD), X-ray photoelectron spectroscopy (XPS) and Fourier transform infrared spectrometer (FTIR) were used to characterize the morphology, elemental composition, crystal structure, function groups of sludge produced by SAEF. UV-visible spectroscopy (UV) and liquid chromatograph-mass spectrometer (LC-MS/MS) were employed to determine the concentration of organic matter, middle products of decomposed organics in the SAEF process, respectively. The results showed that the removal rates of TC, chemical oxygen demand (COD), electric energy consumption (EEC) and the amount of produced sludge (Ws) are 94.87%, 82.42%, 1.383 kWh⋅m⁻³ and 0.1833 kg⋅m⁻³ by SAEF, respectively, under the optimal conditions (pH = 3.0, conductivity (κ) = 1075 μS⋅cm⁻¹, current density (j) = 0.694 mA⋅cm⁻², initial c (TC) = 100 mg·dm⁻³, c [30%H₂O₂] = 1.17 cm³⋅dm⁻³, frequency (f) = 50 Hz, t = 120 min). Compared with pure direct electro-Fenton (DEF) or sinusoidal alternating current coagulation (SACC), SAEF was a highly effective method with low-cost for the treatment of TC wastewater. It was found that the conjugated structure of TC was destroyed to generate intermediate products, and then most of them was gradually mineralized into inorganic materials in the SAEF process.

1. Introduction

Human’s increasingly frequent industrial and agricultural activities led to the production of a large amount of wastewater which contains pesticide residues, heavy metals, antibiotics and other pollutants. If it is discharged directly without any technologies, it results in great pollution to the ecological environment [1,2,3,4,5,6,7,8,9]. Especially, with the increasing abuse of antibiotics, the harm caused by antibiotics to human sustainable development and ecological environment is becoming more and more obvious [10,11]. Tetracycline (TC) is a widely used, great-quality and inexpensive antibiotic, which is commonly used in human health care and fodder. However, because the TC was incompletely absorbed by the gut and decomposed, a great quantity of TC were discharged into the environment in the form of parent compounds, which are not easy to be biodegraded and accumulated in the environment, causing serious pollution to soil, surface water and groundwater. In recent years, TC can be detected in groundwater, and the concentration of TC detected in feces was 20 mg·dm⁻³ [7]. Furthermore, the concentration of TC in the wastewater produced by some pharmaceutical factories can be 400 mg·dm⁻³ [12]. Even after these pharmaceutical factories treat the wastewater, the discharged water can still be detected with a certain concentration of antibiotics [7]. Antibiotics entering the environment will not only lead to chemical pollution, but also induce microbial resistance, accelerate the diffusion and spread in the ecological environment. These resistant microorganisms affect human health through the food chain [13,14]. Therefore, how to efficiently remove TC from wastewater is a research hotspot in the field of water treatment.

The technics of treating TC wastewater mainly include biological treatment [15,16,17], physical-chemical treatment [18,19,20] and their combined processes [21,22,23]. Biotechnology is limited by the growth process of microorganisms, and there is a risk of releasing resistant bacteria and resistant genes into the environment; physical-chemical treatment technology contains advanced oxidation, coagulation, membrane separation and adsorption. However, there are some problems such as high energy consumption, low efficiency, complex process, secondary pollution and so on. Electrocoagulation technique has the advantages of simple equipment, convenient operation, wide application range and high treatment efficiency [24,25,26,27], which is expected to become an effective method for treating TC wastewater. Baran et al. [28] studied that the removal of veterinary antibiotics from wastewater by electrocoagulation. Electrocoagulation is adopted to remove TC with higher removal efficiency. It has been reported that the direct current (DC) was mostly used in electrocoagulation method [29,30,31]. However, direct current coagulation (DCC) process also faced some bottlenecks, mainly in the aspects of electrode passivation and large power consumption. Alternating current (AC) can realize current cycled reversal, which has the advantages of reducing concentration polarization, delaying passivation process, decreasing anode loss and realizing the electrode completely irreversible reaction process. However, there are few literatures about sinusoidal alternating current coagulation (SACC). Sinusoidal waveforms can be applied to the electrocoagulation process with a simple transformer and frequency adjustment. Sinusoidal AC is most easily generated from inexpensive equipment. If the AC frequency of 50 Hz of mains power is sufficient, it can be applied to the electrocoagulation process with a simple transformer.

Based on the previous work [32,33,34,35,36], this research is to analyze the effect of current density, initial pH value, conductivity and H₂O₂ dosage on TC wastewater treatment by sinusoidal alternating electro-Fenton (SAEF). The mechanism of SAEF treating TC wastewater was studied by SEM, FT-IR, XRD, XPS, UV-vis and LC-MS/MS [37,38].

2. Materials and Methods

2.1. Experimental Device

A schematic diagram of TC removal system by SAEF is presented in Figure 1. The reactor (effective volume = 4 dm³) was mainly made of polymethyl methacrylate (PMMA). 6 parallel iron plates were placed vertically inside the reactor, and the gap between the two electrodes was 10 mm. The AC power was used as the power supply which was connected to those iron plates. A certain amount of H₂O₂ solution (mass fraction: 30%) was added into the reactor before the reaction.

2.2. Chemicals and Regents

All chemicals were not further purified prior to the experiment. The Q235B steel plate (100 mm × 300 mm × 2 mm, C ≤ 0.22%, Mn ≤ 1.4%, Si ≤ 0.35%, S ≤ 0.045%, p ≤ 0.045%) were supplied by Aerospace Kaitian Environmental Technology Co., Ltd., Changsha, China. Tetracycline hydrochloride (C₂₂H₂₅ClN₂O₈), hydrogen peroxide (H₂O₂, 30% by mass), sodium chloride (NaCl), sodium hydroxide (NaOH), hydrochloric acid (HCl), ammonium molybdate (H₈MoN₂O₄), potassium dichromate, ferrous sulfate (FeSO₄·7H₂O), o-phenanthroline, concentrated sulfuric acid, silver sulfate powder, mercury sulfate powder, ascorbic acid, potassium persulfate, as analytical grade, were purchased from Pengjin reagent distributor and Sinopharm Group, Changsha, China. The self-made ultra-pure water (18.25 ΜΩ·cm) was used in the experiment.

The instruments: AC power supply (CHP-500VA, Chuanghui Co., Ltd., Shenzhen, China); DC power supply (MS305D, Maisheng Co., Ltd., Guangzhou, China); pH meter (PXSJ-216F, Leici Co., Ltd., Shanghai, China); conductivity meter (DDSJ-308F, Thunder Co., Ltd., Shanghai, China); vertical pressure steam sterilization pot (YXQ-LS-50A, Boxun Co., Ltd., Shanghai, China); UV-visible spectrophotometer (TU-1901, Puxi Co., Ltd., Beijing, China); chemical oxygen demand (COD) rapid digestion unit (575-1, Leici Co., Ltd., Shanghai, China); electronic balance (BSA224S, sartorius Co., Ltd., Shanghai, China); vacuum drying oven (DZF-6020AB, Zhongxing Weiye Co., Ltd., Beijing, China); filtration device (RZK-B01, Yiyang Runze Co., Ltd., Beijing, China).

2.3. Experimental Method

Before the experiment, the iron plate was polished with waterproof abrasive paper (120 mesh), and then the surface rust was removed by washing with 10% dilute hydrochloric acid. The next step was to clean the surface of iron plates with distilled water and put them into the reactor immediately; Samples were filtered by 0.45 μm microporous membranes before being analyzed by UV-visible spectrophotometer and LC-MS/MS. Sludge samples were obtained after filtration by qualitative filter paper (diameter = 9 cm) and vacuum drying.

A stock solution was prepared by dissolving an appropriate amount of TC in the ultrapure water (3 dm³, 100 mg·dm⁻³). The pH value of the solution was adjusted with NaOH and HCl. The addition of NaCl was used to adjust the conductivity of the solution. Experimental parameters such as current density, conductivity and initial pH value were adjusted to a certain value. The supernatant should be filtered by 0.45 μm filter membrane, and the concentrations of TC and COD in the filtrate were determined.

2.4. Measurement

The concentrations of TC in the solution were quantitatively determined by UV-visible spectrophotometer at the wavelength of 357 nm. The concentrations of COD in water were determined by the fast digestion spectrophotometry. TC conversion efficiency (C_TC) and COD removal rate (R_COD) were calculated by Equations (1) and (2), respectively:

$$C_{\mathrm{TC}}=\frac{T_0-T_{\mathrm{t}}}{T_0}\times 100\%$$

(1)

$$R_{\mathrm{COD}}=\frac{C_0-C_{\mathrm{t}}}{C_0}\times 100\%$$

(2)

where, T₀ is the initial TC concentration of TC simulated wastewater raw liquid, mg·dm⁻³; T_t is the TC concentration at time t, mg·dm⁻³; C₀ is the initial COD value of TC simulated wastewater, mg·dm⁻³; C_t is the COD value at time t, mg·dm⁻³.

The pH value and conductivity of solution were measured by PHSJ-4F pH meter and DDSJ-308F conductivity meter, respectively. The surface morphology and element content of the flocs were characterized by field emission environmental scanning electron microscopy (SEM, JSM IT-500, Tokyo, Japan). MAGNA-IR560 spectrometer (Nicolet, Green Bay, WI, USA) was utilized to obtain the Fourier transform infrared (FT-IR) spectra at room temperature by the KBr laminating method. In addition, the composition and structure of the precipitate were analyzed by X-ray diffraction (XRD, Shimadzu 6100, Kyoto, Japan) with a copper target (Kα, λ = 0.154178 nm) at 50 kV and scanning speed of 2°⋅min⁻¹. The degradation products of the samples were determined by liquid chromatography-tandem mass spectrometer (LC-MSMS, Agilent 6470, Santa Clara, CA, USA) under the following conditions: injection volume: 2000 μL; chromatographic column: C18 column, 2.1 × 100 mm, 1.8 μm; the mobile phase was 10 mmol⋅dm⁻³ trifluoroacetic acid-methanol; ion source: electrospray ion source (AJS ESI); positive ion mode; ms scanning range: 200–800 m/z.

3. Results and Discussion

3.1. Optimization of Process Conditions

3.1.1. Effect of Initial pH Value

According to Figure 2, when the initial pH increased from 7 to 9, the TC conversion efficiency decreased from 77.66% to 22.95% at 30 min. As alkaline conditions were more likely to convert iron ions into precipitation, thus inhibiting the generation of ·OH and limiting the oxidation reaction [39]. When the initial pH value was 1, the TC conversion efficiency was also in a low level, because the excessively low pH led to dehydrogenation during the reaction, which hindered the Fenton reaction. By contrast, when the pH value was 3–5, the TC conversion efficiency was much higher. Especially, when the pH was 3, the TC conversion efficiency reached the maximum value.

3.1.2. Effect of Conductivity

The conductivity is an important factor affecting the effect of electrochemical reaction, electrode passivation and electrolytic energy consumption. In the SAEF process, increasing of electrical conductivity improved the electrolytic efficiency of the reaction and accelerated the formation of Fe²⁺ which reacted with H₂O₂.

The effect of different conductivity (0.3, 0.6, 0.9, 1.2 and 1.5 g NaCl were added, respectively) on the TC conversion efficiency were determined. According to Figure 3, when the conductivity increased from 511 μS·cm⁻¹ to 1410 μS·cm⁻¹, the TC conversion efficiency first experienced an upward trend and then decreased. Increasing the concentration of electrolyte in the solution could improve the conductivity of the solution, resulting in the enhancement of both the rate of pollutant migration from the solution to the plate surface and the rate of electron transfer in the reaction system, resulting to accelerate the degradation of TC. In addition, rising the conductivity of the solution could help to reduce the voltage and the hydrogen evolution side reaction. However, when the concentration of NaCl further increased, the TC conversion efficiency decreased. Excessive NaCl might lead to excessive adsorption of salt ions on the electrode plate. Excessive NaCl film on the electrode could reduce the reactive sites and hinder the formation of H₂O₂ and ·OH on the electrode plate. Meanwhile, it might also lead to the formation of Cl₂ and HClO on the electrode plate, which had lower oxidation capacity compared with ·OH. When the electrical conductivity was 1075 μS·cm⁻¹, the TC conversion efficiency reached 91.95% at 30 min, which was the maximum of TC conversion efficiency.

3.1.3. Effect of Current Density

In the SAEF process, increasing current density benefited the electrode reaction. However, it also easily led to electrode polarization, electrode passivation and higher energy consumption. Appropriate current density is the key to TC oxidative degradation.

According to Figure 4, when the current density increased from 0.231 mA·cm⁻² to 0.694 mA·cm⁻², TC conversion efficiency increased from 86.96% to 91.95%. Enhancement of current density could accelerate mass transfer and TC molecule movement towards the electrode. Furthermore, increasing current density could promote electron transfer velocity in redox reactions, thereby promoting the degradation rate of TC. However, when the current density increased to 0.926 mA·cm⁻², the oxidative degradation effect of TC decreased. It might be caused by the rising of voltage of the system, thereby resulting in the occurrence of oxygen or hydrogen gas evolution, and other side reactions at alternating anodes or alternating cathodes, respectively. Excessive current density impeded the generation, accumulation and transformation of H₂O₂, thus the degradation rate was inhibited.

In addition, excessive current density costs much more energy. According to Faraday’s Law, the calculation formulas of electrode plate consumption and electric energy consumption are as follows [40]:

$$m_{\mathrm{Fe}}=\frac{MIt}{1000VzF}$$

(3)

$$W_{\mathrm{e}}=\frac{UIt}{\mathrm{3,600,000}V}$$

(4)

where: m_Fe is plate consumption, kg·m⁻³; W_e is electric energy consumption, kWh·m⁻³; U is voltage, V; I is the current intensity, A; z is the number of electron transfer of iron electrode, and the value is 2. F is Faraday constant, F = 96,485 C·mol⁻¹; V is the volume of actual reaction, m³; M is the molar mass of iron, g·mol⁻¹; t is the reaction time, s.

According to Equations (3) and (4), with the enhancement of I, m_Fe and W_e increase. Increases of U and time also lead to increases in W_e. Therefore, 0.694 mA·cm⁻² was selected as the optimal current density considering the removal effect and cost.

3.1.4. Effect of H₂O₂ Dosage

The applied dosage of H₂O₂ was important for the performance of SAEF process. Appropriate H₂O₂ dosage is beneficial to the formation of ⋅OH, which is conducive to oxidative degradation of TC. However, excessive H₂O₂ may hinder Fenton reaction.

According to Figure 5, when the H₂O₂ dosage increased from 0.50 cm³⋅dm⁻³ to 1.17 cm³⋅dm⁻³, TC conversion efficiency increased from 76.13% to 91.95%. The more H₂O₂ usage, the more ⋅OH generated, thus improving the reaction rate. However, when the H₂O₂ dosage rose to 1.67 cm³⋅dm⁻³, TC conversion efficiency declined. Since excessive H₂O₂ reacted with ⋅OH to form weak oxidant HO₂ and reduced the amount of ⋅OH in wastewater, which was not conducive to the oxidative degradation of TC. Furthermore, excessive H₂O₂ increased the cost. Therefore, the optimal condition was when the dosage of H₂O₂ is 1.17 cm³⋅dm⁻³.

3.2. SAEF Process Analysis and Comparison

3.2.1. Analysis of Treatment Effect of SAEF Process

Under the optimal conditions, the treatment effect of TC wastewater and the change of pH value with time were investigated, as shown in Figure 6.

According to Figure 6a, the TC conversion efficiency reached 88.73% after 5 min and maintained stable. COD concentration generally showed a downward trend. COD is an important target to measure the pollution degree of water and could be used to illustrate the degree of complete mineralization of organic matter [41]. When the reaction lasted for 120 min, the TC conversion efficiency and COD removal rate were 94.87% and 82.42%, respectively. It implies that 82.42% TC was completely mineralized, and 12.45% TC was converted into other intermediate organic products.

During the SAEF progress from 0–2 h, the pH value was changed from 3 to 3.35, as shown in Figure 6b. Generally, the optimal pH value range lies at 3–5 in Fenton reaction [42], but the optimal operating pH value range is 5–9 in electrocoagulation [43]. Under the condition of pH 3–5, Fenton reaction was the dominate reaction, but the removal of organic matter was accompanied by subordinate coagulation and adsorption of iron ion hydrolysate. The hydrolysis product of iron ion was greatly affected by pH value. At low pH values, the main forms of iron ions were complex ions with high charge and low polymerization, such as Fe(OH)²⁺ and Fe(OH)³⁺, and the mechanism of coagulation was mainly charge neutrality. Low pH was not conducive to the generation of polynuclear hydroxyl complexes and hydroxides, but it was beneficial to reduce electrode passivation and dissolve oxide film. Furthermore, low pH could promote Fenton reaction, which was conducive to oxidative degradation of TC. When pH value increased, iron ion polyhydroxy complex would be rapidly produced, and the pollutants in the solution would be removed by surface complexation and electrostatic adsorption.

In the SAEF reaction system, under the action of 50 Hz alternating electric field, the iron electrodes slowly producing Fe²⁺ alternately. At the beginning of the reaction, H₂O₂ and Fe²⁺ produced a lot of ·OH, which attacked the conjugated double bond structures of TC molecules. Thus, high TC conversion efficiency was achieved. Afterwards, TC conversion efficiency did not increase significantly, while COD removal rate continued to rise, which indicates that the intermediate products decomposed in TC molecule were still degraded by oxidation by Fenton reaction. While Fe²⁺ is oxidized into Fe³⁺ by Fenton reaction and anodic oxidation at electrode surface.

3.2.2. Comparison of Treatment Effects among SAEF, DEF and SACC

Under the optimal conditions, treatment effects of SAEF, direct electro-Fenton (DEF) and SACC (to achieve better electrical flocculation effect, the initial pH value of SACC was set at 7) were compared. The reaction time was 120 min, and the results are shown in Figure 7.

By comparing the treatment effects of SAEF and DEF, we find that the TC conversion efficiency of SAEF is 5.07% lower than that of DEF, the COD removal rate of SAEF and DEF was nearly the same, but the sludges output and energy consumption of DEF were 3.23 times and 1.15 times of SAEF, respectively. In the electro-Fenton reaction system, Fe²⁺ was a necessary condition for catalyzing H₂O₂ to produce ⋅OH [44]. Compared with DC, the dissolution rate of Fe²⁺ was slower under the action of 50 Hz AC, resulting in a lower concentration of Fe²⁺ in the solution. The amount of ⋅OH generated by H₂O₂ and Fe²⁺ were smaller. Therefore, the oxidation degradation effect of TC in DEF was better than that in SAEF. Under the condition of DC, the dissolution rate of Fe²⁺ and the formation rate of ⋅OH were fast, and the higher voltage led to strong anodic oxidation. However, excessive Fe²⁺ dissolution led to a great increase of flocs. Higher voltage resulted in higher power consumption.

By comparing the treatment effects of SAEF and SACC, the TC conversion efficiency and COD removal rate of SAEF were 28.05% and 21.83% higher than that of SACC, respectively. The sludges production of SAEF was slightly higher than that of SACC. However, the energy consumption of SAEF was slightly lower than that of SACC. Obviously, the difference of treatment effect between SAEF and SACC was caused by Fenton oxidation. The amount of Fe²⁺ dissolved by SAEF was more than that of SACC under acidic condition, so the amount of sludge produced by SAEF increased relatively. Comparing with SACC, SAEF had more redox reactions, which might be related to the generation of ammonia, nitrate, nitrite and other inorganic ions in the degradation process of TC. SAEF showed relatively high conductivity and low voltage, thereby resulted in low energy consumption.

It was found that the sedimentation rate of the flocs generated by AC power supply was much lower than that generated by DC power supply during the standing process, which indicated that the sludges generated by AC power supply had better dispersion and smaller particle size.

3.3. Characterization Analysis

3.3.1. The SEM Analysis

Figure 8a,b shows the SEM analysis of the flocs collected after the reaction of SAEF and DEF. According to Figure 8a, the flocs generated by SAEF showed a granular structure and good dispersion. The alternating direction of applied electric field in the AC field makes agglomeration difficult. According to Figure 8b, the flocs produced by DEF showed poor dispersibility and tend to be clustered together in a block structure. As the reaction rate was fast, the rate of ion dissolution and ion mass transfer in the iron plate was faster under the action of DC. Compared with the AC power supply, the DC power supply could generate more flocs in the same period. The results showed that the dispersion of SAEF flocs is better than that of DEF flocs.

3.3.2. The FT-IR Analysis

Figure 9a is the FT-IR spectroscopic analysis of flocs collected after the reaction of SAEF and DEF. The absorption peaks at 3346 cm⁻¹ and 3364 cm⁻¹ were caused by O-H stretching vibration in H-O-H, and the peak at 1645 cm⁻¹ and 1650 cm⁻¹ might be caused by the skeleton vibration of benzene ring or H-O-H bending vibration [45]. The H-O-H peak was caused by O-H and adsorbed water in the sample. The absorption peak at 1403 cm⁻¹ might be caused by the adsorption of some functional groups in TC by floc, such as C-N bending vibration or -CH₂- bending vibration. The absorption peaks at 486 cm⁻¹, 490 cm⁻¹ and 675 cm⁻¹ were characteristic absorption peaks of Fe-O bond, which were caused by the stretching vibration and lattice vibration of Fe-O in iron oxide. FT-IR analysis showed that the flocs might contain Undegraded TC or its degradation intermediates.

3.3.3. The XRD Analysis

Figure 9b demonstrates the XRD analysis of SAEF and DEF flocs. During the reaction, the Fe²⁺ was produced by electrochemical oxidation. Then the Fe²⁺ was further oxidized into the corresponding flocs through a series of physical and chemical reactions. The crystalline of the flocs created by SAEF is not obvious, which might be caused by the small size of flocs particles. The XRD pattern of flocs obtained by DEF showed two obvious characteristic diffraction peaks, which was coincide with FeOOH (JCPDS NO:13-0087). 2θ = 35.2° and 2θ = 63.1° were the crystal surface of FeOOH, which indicates that the flocs contain FeOOH.

3.3.4. The XPS Analysis

To further study the mechanism of TC degradation during the SAEF, XPS analysis was carried out. According to Figure 10a, Fe, C, N and O elements are observed in the flocs. According to the Fe 2p spectrogram in Figure 10b, two major asymmetric peaks are located at 710.6 eV (satellite peak at 718.1 eV) and 723.9 eV, which belong to Fe 2p3/2 and Fe 2p1/2, respectively. For Fe 2p3/2, the peak at 710.6 eV is considered as Fe(II), while the other peak at 714.4 eV is considered as Fe(III). For the spectra of C1s (Figure 10c), the three binding energy peaks at 284.80 eV, 286.70 eV and 288.30 eV belong to C=C, C-C and C=O groups, respectively. For the spectra of N1s (Figure 10d), the binding energy peak at about 399.60 eV belongs to the C-N bond of organic matter. For the spectrum of O1s, as shown in Figure 10e, there is an obvious peak at 531.08 eV, which belongs to C=O double bond. The 529.7 eV is O-H bond. The results of XPS analysis showed that TC and its intermediates were contained in the flocs.

3.3.5. The UV-Vis Analysis

The structure information of conjugate system reflected in UV-vis spectra. TC showed characteristic absorption in ultraviolet and visible zone. Figure 11 illustrates UV-vis spectral of water samples in SAEF and SACC systems at different reaction times.

In the SAEF process, there were two obvious absorption peaks of TC at 274 nm and 357 nm before the reaction, which corresponded to TC. As the reaction progressed, the peak intensity decreased rapidly at about 357 nm, indicating that TC was removed. Furthermore, The occurrence of bathochromic-shift (around 368 nm) indicated that TC or its intermediate products might have a complex reaction with Fe(III) in the degradation process. No obvious absorption peak at 357 nm after 5 min. Peak intensity at 274 nm decreased gradually, which indicated that TC in the wastewater was effectively degraded and removed. In the SACC system, with the progress of the reaction, the peak intensity decreased rapidly around 357 nm, indicating that TC was continuously degraded, accompanied by a certain degree of red shift to 368 nm. However, peak intensity at 274 nm gradually decreases and the absorption peak showed a hypochromatic shift. On the one hand, the blue shift might be caused by the reduction of conjugation degree, which indicated that the conjugation structure of TC was rapidly destroyed and might be decomposed into smaller molecular fragments. On the other hand, TC might be complexed with Fe(III) and generated stable Fe(III)-TC complex [46].

The characteristic absorption peak at 357 nm reduced obviously after 1 min in the SAEF process, and there was no obvious absorption peak at 357 nm after 3 min. Furthermore, the absorption peak at 274 nm also declined obviously, indicating a fast reaction rate. However, the characteristic absorption peak of 357 nm in the SACC process existed during the whole reaction, indicating the poor removal effect and the low reaction rate. The above result analysis showed that SAEF had better performance than that of SACC, which means the Fenton reaction played a significant role in the whole reaction.

3.3.6. The LC-MS/MS Analysis

The LC-MS/MS total ion chromatograms and detected components are shown in Figure 12 and

Table 1. LC-MS/MS determination of TC electrochemical oxidation intermediates.

Serial Number	m/z	Ion Mass-to-Charge Ratio	Molecular Formula	Structural Formula	Peak Position
1	509	491, 447	C22H24O12N2		10, 11, 13, 14
2	448	430, 413	C20H17O11N		1
3	451	433, 417	C19H14O13		11
4	396	379, 361	C15H9O12N		3, 4, 7, 8, 9, 10, 11, 13, 14
5	367	349	C15H10O11		4
6	253	235	C10H4O8		9,10,13

, respectively. According to the LC-MSMS results of water sample in the SAEF process, a variety of intermediate products were found during TC degradation, including C₂₂H₂₄O₁₂N₂, C₂₀H₁₇O₁₁N, C₁₉H₁₄O₁₃, C₁₅H₉O₁₂N, C₁₅H₁₀O₁₁, C₁₀H₄O₈, etc. In combination with the structure and related chemical principles of TC, C=C, C-N and benzene ring in TC are easily attacked by ·OH during the TC degradation process. Macromolecular intermediates were generated through a series of reactions, which were further oxidized and rate-opening converted into small molecular organic compounds, and finally converted into inorganic substances such as CO₂, H₂O, NO₃⁻ and NH₄⁺.

3.3.7. The Mechanism Analysis

Based on the above discussion, the Figure 13 illustrates the mechanism of SAEF reaction and the process of TC removal, which including Fenton reaction, Fe³⁺ involved oxidation, anodic oxidation and flocculation adsorption, among which Fenton reaction plays a leading role. Specifically: (1) Fenton reaction process. ·OH is generated from the dissolved Fe²⁺ and H₂O₂, and TC molecules and their intermediates are oxidized by ·OH. (2) Fe²⁺ participates in the oxidation process. Fe²⁺ is dissolved by iron electrode, and Fe²⁺ complexes with TC molecule to form Fe(II)-TC complex. Fe²⁺ and Fe(II)-TC complex can oxidize organic matter in their own oxidation process. (3) anodic oxidation process. Part of TC or intermediate products adsorbed on the anode surface were directly oxidized or indirectly oxidized by Cl₂ and HClO generated in the reaction process. (4) flocculation air flotation process. The coagulation and adsorption of iron ion hydrolysate can remove TC or intermediate products to a certain extent. In addition, the H₂ produced by the cathode makes the small density floc float up and separate, which helps to quickly remove the pollutants in the water. TC is first oxidized to form macromolecular intermediates, then transformed into small molecular organic matter through oxidation and ring-opening, and finally transformed into inorganic substances such as CO₂, H₂O, NO₃⁻ and NH₄⁺.

4. Conclusions

In this study, the treatment of TC wastewater by SAEF was reported for the first time, and the conclusions are as follows:

(1) SAEF is a novel process which can effectively remove TC in wastewater. When the current density (j) was 0.694 mA⋅cm⁻², the conductivity (κ) was 1075 μS⋅cm⁻¹, c [30%H₂O₂] = 1.17 cm³⋅dm⁻³, initial pH was 3.0, and the reaction time (t) was 120 min, TC and COD removal rates were 94.87% and 82.42%, respectively.

(2) In Fenton reaction of the SAEF process, Fe²⁺ takes part in the processes of oxidation by solution, anodic oxidation and flocculation adsorption, which were contributed to the removal of TC. Fenton reaction process played a leading role. The TC degradation process included that TC was firstly oxidized to generate macromolecular intermediates, then transformed into small molecular organic matter through oxidation and ring-opening, and finally transformed into inorganic substances such as CO₂, H₂O, NO₃⁻ and NH₄⁺.