Enhanced Fenton-like Degradation of Trichloroethylene by Hydrogen Peroxide Activated with Nanoscale Zero Valent Iron Loaded on Biochar

Abstract

Composite of nanoscale Zero Valent Iron (nZVI) loaded on Biochar (BC) was prepared and characterized as hydrogen peroxide (H₂O₂) activator for the degradation of trichloroethylene (TCE). nZVI is homogeneously loaded on lamellarly structured BC surfaces to form nZVI/BC with specific surface area (S_BET) of 184.91 m² g⁻¹, which can efficiently activate H₂O₂ to achieve TCE degradation efficiency of 98.9% with TOC removal of 78.2% within 30 min under the conditions of 0.10 mmol L⁻¹ TCE, 1.13 g L⁻¹ nZVI/BC and 1.50 mmol L⁻¹ H₂O₂. Test results from the Electron Spin Resonance (ESR) measurement and coumarin based fluorescent probe technology indicated that ∙OH radicals were the dominant species responsible for the degradation of TCE within the nZVI/BC-H₂O₂ system. Activation mechanism of the redox action of Fe²⁺/Fe³⁺ generated under both aerobic and anaerobic conditions from nZVI and single electron transfer process from BC surface bound C–OH to H₂O₂ promoted decomposition of H₂O₂ into ∙OH radicals was proposed.

Introduction

As one of the most commonly used chemicals in industry, chlorinated solvents such as trichloroethylene (TCE) has been frequently encountered in subsurface environments as dense non-aqueous phase liquids (DNAPLs) at many industrial sites^1,2. Due to its high toxicity and adverse effects on liver and kidney, TCE has been classified as a potential human carcinogen and listed as a priority pollutant by the United States Environmental Protection Agency (U.S. EPA)³, and the Safe Drinking Water Act in the USA defines the maximum contaminant level of TCE at 5 μg L⁻¹ for drinking water⁴. Therefore, effective remediation and complete mineralization of TCE in aquifers is urgently required to reduce its adverse effects on the environment and human health.

Advanced oxidation processes (AOPs) have been emerged as the most efficient alternative to degrade various organic pollutants for the generation of reactive radicals⁵. Among AOPs, Fenton (i.e., the reaction between Fe²⁺ and H₂O₂) is a powerful oxidant generating the hydroxyl radicals (∙OH, E₀ = 2.80 V), which react with various organic compounds at the near-diffusion controlled rates^6,7, leading to an effective degradation and mineralization of organic pollutants⁸. However, it should to be operated at pH < 3.0 and the generated iron sludge limited the wide application of the homogeneous Fenton process⁹. Heterogeneous Fenton-like activator of zero valent iron (ZVI) was developed, which could be used over a wide pH range to decompose H₂O₂ instead of homogeneous ferrous iron^10,11. In addition, nanoscale zero valent iron (nZVI) could enhance H₂O₂ activation due to the small particle size and high reactivity. Xu et al.¹² reported that the heterogeneous Fenton-like system using nZVI as catalyst was effective for the removal of biocide 4-chloro-3-methyl phenol in the presence of H₂O₂, and the reaction was induced through following reactions in the heterogeneous system of nZVI/H₂O₂:

Though nZVI has performance for H₂O₂ activation, it tends to aggregate into forming microscale particles due to its high surface energy and strong magnetic interaction, leading to the reduced reactivity¹³. To overcome the problem, granular activated carbon¹⁴, bentonite¹⁵, and rectorite¹⁶ were introduced as a support for nZVI to gain better distribution. Biochar (BC) is a promising environmental friendly material pyrolyzed under low oxygen conditions. It possesses large surface area with porous structure and has oxygen containing functional groups¹⁷. Thus, it is anticipated that nZVI loaded uniformly on BC surface to form nZVI/BC composite will effectively prevent the aggregation of nZVI with significantly enhanced performance.

In this study, the composite of nZVI loaded on BC sheets was synthesized and characterized as H₂O₂ activator for the degradation of TCE. The presence of C−OH groups on BC surface could be used to activate H₂O₂ through electron transfer process¹⁸. Thus, both the dispersive nature and H₂O₂ activator of BC will be simultaneously achieved for nZVI/BC. The present work aims to (1) synthesize a novel composite of nZVI/BC, where nZVI was loaded on BC surface uniformly and the aggregation of nZVI was prevented effectively, (2) characterize nZVI/BC activation ability for H₂O₂ to degrade TCE in aqueous solution, and (3) explore the activation mechanisms of H₂O₂ in the presence of nZVI/BC.

Results and Discussion

Characterization of nZVI/BC

SEM analyses were firstly conducted to observe the morphologies of the prepared nZVI, BC and nZVI/BC, respectively. As illustrated in Fig. 1a, nZVI was spherical with diameters of about 30 nm, and the agglomeration of nZVI was observed due to the nanometer effect and magnetic properties of nZVI. In addition, lamellarly structured BC of rough surface morphologies was obtained, and nZVI was homogeneously loaded on BC surface from the SEM image of nZVI/BC composite. XRD was also conducted, and the data were shown in Fig. 1d. It can be seen that the XRD pattern of nZVI revealed a highly crystalline and single phase structure by diffraction peaks at 45.0° (JCPD 01-087-0721)¹⁹. The crystallite size of nZVI was estimated to be 29.7 nm derived from the Debye-Sherrer equation (D = Kλ/(βcosθ), where K is the Sherrer constant (0.89), λ is the X-ray wavelength (0.15418 nm), β is the full peak width at half maximum and θ is the Bragg diffraction angle), which was in accordance with the SEM image. The broad reflection peak of BC in XRD indicated the amorphous BC, suggesting that nZVI was successfully loaded on BC surface from XRD spectrum of nZVI/BC (Fig. 1d).

Figure 1

SEM images of (a) nZVI, (b) BC, (c) nZVI/BC, (d) XRD patterns of nZVI, BC and nZVI/BC, (e) nitrogen adsorption-desorption isotherms of nZVI, BC and nZVI/BC, and (f) FT-IR spectra of nZVI, BC and nZVI/BC.

The BET surface areas were measured by using N₂ adsorption method as shown in Fig. 1e. The S_BET values were calculated according to , where V is the volume of nitrogen adsorbed per gram, V_m is the monolayer capacity and C is related to the heat of adsorption. From the results, the S_BET value of bare nZVI was 26.61 m² g⁻¹, and was increased to 184.91 m² g⁻¹ for the nZVI/BC composite after nZVI was loaded onto BC surface (S_BET value of 205.35 m² g⁻¹). In the FTIR spectrum, the band at about 3400 cm⁻¹ was belonged to the vibration of hydroxyl groups (−OH). The signal at 1590 cm⁻¹ was ascribed to C=O stretching vibration, the peak at 1100 cm⁻¹ was assigned to the C−O groups, and the absorption peak at 802 cm⁻¹ was corresponded to the vibration of aromatic C−H¹⁷. The weak adsorption in nZVI/BC spectrum at 561 cm⁻¹ was observed in the Fig. 1f, indicating the Fe−O bond formed between BC and nZVI²⁰.

Heterogeneous fenton-like degradation of TCE in nZVI/BC-H₂O₂ system

The performances of TCE degradation by H₂O₂ activated with nZVI, BC and nZVI/BC were investigated and presented in Fig. 2a. The control test suggested the TCE (0.10 mmol L⁻¹) loss was less than 2% due to the volatilization during the experimental period under all the tested conditions. With the effect of 1.50 mmol L⁻¹ H₂O₂, TCE was hardly degraded in the absence of any activators within 30 min, but its degradation efficiencies were enhanced to 39.1%, 6.5% and 98.9% with H₂O₂ in the presence of nZVI, BC and nZVI/BC, respectively. Under all the tested conditions, the TCE degradation kinetics approximately followed pseudo-first-order reaction of ln (c₀/c_t) = kt + b, where c₀ and c_t are the TCE concentrations at initial time (t = 0) and reaction time (t = t), k is the apparent reaction rate constant (min⁻¹), t is reaction time (min), and b is a constant. As illustrated in Fig. 2b by adding 0.19 g L⁻¹ nZVI, the apparent reaction rate constant of TCE was 0.0128 min⁻¹, and the relative small k value of TCE degradation might be due to the aggregation of prepared nZVI with a small S_BET value. The k value of 0.0002 min⁻¹ for TCE degradation was observed when 0.94 g L⁻¹ BC was added, indicating the weak activation ability of BC for H₂O₂, which was in accordance with the results reported by previous studies^21,22. Loading nZVI onto BC significantly increased k values almost by 11 times from 0.0128 to 0.136 min⁻¹ from Fig. 2a, demonstrating that nZVI/BC was more efficient than nZVI for H₂O₂ activation. nZVI were distributed on the BC surface homogeneously from SEM image, and nZVI/BC has higher S_BET value than that of the raw nZVI. The increased surface area of nZVI/BC enhanced the amount of H₂O₂ activation sites and TCE adsorption, which might significantly lead to the increase in the H₂O₂ activation ability and TCE degradation rate constant.

Figure 2

(a) Kinetic data of TCE in the systems of H₂O₂, nZVI-H₂O₂, BC-H₂O₂ and nZVI/BC-H₂O₂, and the inset was the plot of ln(c/c₀) vs. t within the four mentioned systems, effects of (b) nZVI/BC dosage, (c) H₂O₂ concentration and (d) solution pH on apparent reaction rate constant of TCE degradation. Reaction conditions: the concentration of nZVI itself or in nZVI/BC composite was 0.19 g L⁻¹, the dosage of nZVI/BC was 1.13 g L⁻¹, the dosage of BC was 0.94 g L⁻¹, the concentration of H₂O₂ was 1.50 mmol L⁻¹, the concentration of TCE was 0.10 mmol L⁻¹, and the initial pH was 6.2.

As shown in Fig. 2b and c, effects of nZVI/BC dosage and initial H₂O₂ concentration were investigated. When nZVI/BC dosage was increased from 0.372 to 1.13 g L⁻¹, the k value for TCE degradation was almost linearly increased from 0.022 to 0.136 min⁻¹. The increased nZVI/BC dosage provided a large amount of H₂O₂ active sites, and thus more ∙OH radicals generated²³. However, the k value was decreased to 0.101 min⁻¹ when the dosage of nZVI/BC was further increased to 1.13 g L⁻¹. The decreased k value may be ascribed to the decrease of ∙OH radicals for TCE degradation due to the scavenging effect by excessive Fe²⁺ species (generated from Eqs 1 and 2) in according to Eq. 4 ²⁴.

In consideration of the effect of H₂O₂ concentration, the k values were increased quickly from 0.033 min⁻¹ to 0.136 min⁻¹ as H₂O₂ concentrations were increased from 0.33 mmol L⁻¹ to 1.50 mmol L⁻¹. H₂O₂ is the precursor for ∙OH generation, relative high H₂O₂ concentration induced more ∙OH radicals accounted for TCE degradation, and hence the increased k value was obtained. When the concentration of H₂O₂ was beyond 1.50 mmol L⁻¹, the k value was decreased due to the reaction between ∙OH and excessive H₂O₂ (Eq. 5)^8,25. Therefore, the nZVI/BC dosage and the initial H₂O₂ concentration were fixed to be 1.13 g L⁻¹ and 1.50 mmol L⁻¹ respectively for the degradation of 0.10 mmol L⁻¹ TCE.

Figure 2d showed the effect of the initial solution pH on TCE degradation in the presence of nZVI/BC and H₂O₂. The ability of nZVI/BC to activate H₂O₂ was decreased with the increase of the initial solution pH. However, the k value of 0.059 min⁻¹ was observed when the solution pH was as high as 10.0, which indicated that the nZVI/BC-H₂O₂ system could be effective even in alkaline pH conditions with no adjustment of the pH value being needed for the effective TCE degradation.

TCE mineralization and the stoichiometry efficiency of utilization of H₂O₂

TOC analyzer was used to evaluate the mineralization of TCE in the presence of H₂O₂ activated by nZVI/BC. The degradation of TCE was monitored, and the stoichiometry efficiency of utilization of H₂O₂ was calculated. In accordance with Eq. 6, three moles H₂O₂ will be consumed to obtain the complete mineralization of one mole TCE in theory. The stoichiometry efficiency of utilization of H₂O₂ (η) was the ratio of the amount of H₂O₂ consumed for the TCE degradation (Δ[H₂O₂]_degradation) to the total amount of the H₂O₂ decomposed in the reaction (Δ[H₂O₂]_{decomposition}) in accordance with Eq. 7 ²⁶. The value of Δ[H₂O₂]_degradation was calculated by measuring the TOC change in the TCE solution. The amount of Δ[H₂O₂]_{decomposition} at different reaction time was measured as shown in Fig. 3a.

Figure 3

(a) H₂O₂ consumption and (b) TOC removal in nZVI-H₂O₂, BC-H₂O₂ and nZVI/BC-H₂O₂ systems. Reaction conditions: Reaction conditions: the concentration of nZVI itself or in nZVI/BC composite was 0.19 g L⁻¹, the concentration of nZVI/BC was 1.13 g L⁻¹, the dosage of BC was 0.94 g L⁻¹, the concentration of H₂O₂ was 1.50 mmol L⁻¹, the concentration of TCE was 0.10 mmol L⁻¹, and the initial pH was 6.2.

As shown in Fig. 3b in the presence of 1.50 mmol L⁻¹ H₂O₂, the TOC removal after 30 min was 5.1%, 32.6% and 78.2% with the addition of BC, nZVI and nZVI/BC, corresponding to the TCE degradation efficiencies of 6.5%, 39.0% and 98.9% respectively from GC results (Fig. 2a), respectively. The total amount of the H₂O₂ decomposed in the reaction in BC-H₂O₂-TCE, nZVI-H₂O₂-TCE and nZVI/BC-H₂O₂-TCE systems were 1.76, 12.91 and 15.84 μmol respectively. Therefore, the calculated efficiencies for the utilization of H₂O₂ was 48.7%, 33.3% and 65.2% in the above three systems, respectively. The TOC removal and the efficiency for the utilization of H₂O₂ in nZVI/BC-H₂O₂-TCE system were consistently higher than those for BC-H₂O₂-TCE and nZVI-H₂O₂-TCE systems indicating the superior activation property of nZVI/BC for H₂O₂ after nZVI was loaded on BC surface.

Determination of free radicals

Based on previous studies^27,28,29,30, reactive oxygen species (ROSs, such as ∙OH radicals and O₂⁻∙/HO₂∙ radicals) were generated from the decomposition of H₂O₂ in homogeneous Fenton reactions at acidic and neutral conditions. In addition, high valent iron (Fe^IV = O) was also proposed in heterogeneous Fenton-like reaction during the activation of H₂O₂ by zero-valent iron under alkaline conditions^31,32. To ascertain the H₂O₂ activation mechanism, the involved ROSs in nZVI-H₂O₂, BC-H₂O₂ and nZVI/BC-H₂O₂ systems were monitored by using both DMPO spin trap ESR measurement and coumarin based fluorescent probe technology.

As shown in Fig. 4a, no signals were observed in the presence of H₂O₂. The measured ESR spectra in the three systems mentioned above illustrated the four-fold characteristic peak with an intensity ratio of 1:2:2:1, which were in accordance with the pattern of typical DMPO-∙OH adduct³³. Possibly due to the weak activation ability of BC and high adsorption of DMPO-∙OH adduct by BC in BC-H₂O₂ system¹⁷, relative low concentration of DMPO-∙OH adduct was detected in aqueous solution.

Figure 4

(a) DMPO spin-trapping ESR spectra of ∙OH radicals and (b) emission spectra intensity of coumarin adduct (excited at 346 nm, and detected at 456 nm) in the systems of H₂O₂, nZVI-H₂O₂, BC-H₂O₂ and nZVI/BC-H₂O₂. Reaction conditions: the concentration of nZVI itself or in nZVI/BC composite was 0.19 g L⁻¹, the concentration of nZVI/BC was 1.13 g L⁻¹, the dosage of BC was 0.94 g L⁻¹, the concentration of H₂O₂ was 1.50 mmol L⁻¹, the concentration of TCE was 0.10 mmol L⁻¹, and the initial pH was 6.2.

The intensity of DMPO-∙OH adduct in the nZVI/BC-H₂O₂ system was much higher than that in the nZVI-H₂O₂ system, indicating more ∙OH radicals being generated. In addition, due to the instability of the O₂∙⁻/HO₂∙ radicals in the solution, six-fold characteristic peak of the O₂∙⁻/HO₂∙ radicals adduct by using dimethyl sulfoxide as solvent was measured³⁴, with no signal being detected (data were not shown). These results suggested that ∙OH radicals were the main ROSs generated from the decomposition of H₂O₂ responsible for TCE degradation. Coumarin based fluorescent probe technology was used to measure the generation of ∙OH radicals. As illustrated in Fig. 4b the fluorescence intensities were increased quickly in the first few minutes in both nZVI/BC-H₂O₂ and nZVI-H₂O₂ systems, indicating the fast generation of ∙OH radicals. The fluorescence intensity in the nZVI/BC-H₂O₂ system was consistently higher than that in the nZVI-H₂O₂ system, hinting the excellent property of nZVI/BC for H₂O₂ activation. The data obtained here was well coincided with ESR results.

Discussion on reaction mechanism

nZVI particles could be oxidized to Fe²⁺ under anaerobic or aerobic conditions, and it is known that homogeneous Fe²⁺ plays a critical role in the activation of H₂O₂ to generate ∙OH radicals^35,36. Thus, the dissolved Fe²⁺ and Fe³⁺ concentrations in the presence of nZVI/BC with and without H₂O₂ were measured after reaction, respectively. As shown in Fig. 5a, the concentrations of Fe²⁺ and Fe³⁺ were 0.73 and 3.15 mg L⁻¹ respectively with only the nZVI/BC in aqueous solution. However, in the nZVI/BC-H₂O₂ system, the concentrations of Fe²⁺ and Fe³⁺ were 7.70 and 45.59 mg L⁻¹ respectively, being much higher than those in the absence of H₂O₂. The results indicated that the oxidation of nZVI to Fe²⁺ was occurred initially, and the dissolved Fe²⁺ directly activated H₂O₂ to generate ∙OH radicals subsequently. The consumption of Fe²⁺ for H₂O₂ activation accelerated nZVI transformation and Fe³⁺ formation in accordance with Eq. 3.

Figure 5

(a) Dissolved iron concentrations in the aqueous solution with and without H₂O₂ and (b) C 1 s XPS spectra of nZVI/BC before and after the reaction. Reaction conditions: the concentration of nZVI itself or in nZVI/BC composite was 0.19 g L⁻¹, the concentration of nZVI/BC was 1.13 g L⁻¹, the dosage of BC was 0.94 g L⁻¹, the concentration of H₂O₂ was 1.50 mmol L⁻¹, the concentration of TCE was 0.10 mmol L⁻¹, and the initial pH was 6.2.

In the BC-H₂O₂ system, the degradation efficiency of TCE was 6.5%, indicating that BC has also acted as an electron-transfer mediator to activate H₂O₂. BC characteristics of porosity, specific surface area, surface inertness and surface functional groups might significantly affect the catalytic activity for H₂O₂ decomposition^37,38. XPS spectra of nZVI/BC were measured to better understand the roles of BC in the activation of H₂O₂ before and after the reaction. As illustrated in Fig. 5b, the peaks of C (1 s) at 284.5, 286.5 and 289.0 eV were attributed to C–C, C–OH and COOH, respectively. From the spectra, the proportion of C–C, C–OH and COOH peaks were 71.8%, 21.7% and 6.5% respectively for fresh nZVI/BC before the reaction. However, the proportion of the three peaks mentioned above were 70.0%, 17.3% and 12.7% after the H₂O₂ activation. The decrease in surface bound C–OH proportion of BC after the reaction suggested that BC might act as an electron transfer medium participated in the H₂O₂ activation^39,40, and the increase in COOH was due to the reaction between C–OH and H₂O₂ in accordance with Eq. 8 ¹⁸. By releasing organic radicals of CO∙ from C–OH through single electron transfer process, ∙OH radicals were generated (Eq. 9)⁴¹.

As derived from the above discussions, the activation mechanism of H₂O₂ in the presence of nZVI/BC was proposed in Fig. 6. Firstly, nZVI particles were oxidized to Fe²⁺, and the redox reaction of Fe²⁺/Fe³⁺ was accounted for the generation of ∙OH radicals. Secondly, as an electron transfer mediator to H₂O₂, BC surface bound C–OH decomposed H₂O₂ into ∙OH radicals by releasing CO∙ radicals. Once ∙OH radicals having been produced, it would rapidly react with TCE. GC-MS was utilized to monitor the process of TCE degradation, however, no intermediate products was detected except for the undegraded TCE. Though TCE would be transformed into low molecule weight organic acids with the effect of ∙OH radicals initially, only CO₂ and Cl⁻ were measured during the oxidative process in the nZVI/BC-H₂O₂-TCE system.

Figure 6

A proposed mechanism for TCE degradation by the nZVI/BC activation of H₂O₂.

Conclusions

The nZVI/BC composite was successfully synthesized and characterized as an efficient H₂O₂ activator for the degradation of TCE. nZVI loaded on lamellarly structured BC surface prevented its agglomeration behaviour, whcih significantly enhanced the generation of ∙OH radicals. The redox effect of Fe²⁺/Fe³⁺ and the single electron transfer process from BC surface bound C–OH to H₂O₂ were accounted for the promoted generation of ∙OH radicals, leading to rapid TCE degradation. The enhanced Fenton-like activation of H₂O₂ using nZVI/BC presents the great potential for TCE degradation in aqueous solution.

Materials and Methods

Chemicals and materials

FeSO₄ ∙ 7H₂O, NaBH₄, H₂O₂ (30%, w/w), 5,5-Dimethyl-1-pyrrolidine N-oxide (DMPO) and trichloroethylene (TCE, 99.0%) were purchased from Sinopharm Chemical Reagent Co. Ltd. HCl and NaOH were obtained from Shanghai Lingfeng Chemical Reagent Co. Ltd. All chemicals used in this work were used as received without further purification.

Synthesis of nZVI/BC composite

Biochars were prepared by the pyrolysis of rice hull collected locally. Firstly, the rice hull was washed with ultrapure water and dried in oven at 80 °C. Secondly, the dried rice hull was pyrolyzed in muffle furnace under oxygen limited condition at a temperature of 350 °C for 6 h. Finally, the BC were obtained after treating with 1.0 mol L⁻¹ HCl and washed with ultrapure water.

For the synthesis of nZVI/BC composite, 3.78 g biochar was dispersed in 250 mL oxygen free ultrapure water. Then, 0.0135 mol FeSO₄ ∙ 7H₂O was added at pH 5.0. With mechanical stirring and ultrasonic, nZVI was formed and loaded on the surface of BC by addition of 100 mL 0.27 mol L⁻¹ NaBH₄ at a velocity of 20 mL min⁻¹. Following 2 h reaction, the nZVI/BC composite were separated and washed with deoxygenized ultrapure water, and finally vacuum dried. The preparation of nZVI was described as Eq. 10, and the schematic for the preparation of nZVI/BC composite was shown in Fig. 7.

Figure 7

The schematic for the preparation of nZVI/BC composite.

Characterization

X-ray diffraction (XRD, X’TRA, Swiss) analysis was conducted to determine the crystal structure and crystallinity of the prepared composites using Cu K_α radiation with 2θ collection range from 10° to 80°. X-ray photoelectron spectra (XPS) were recorded on Axis Ultra spectrometer (Kratos) using Al Kα radiation excitation source. The infrared spectrum was recorded on Fourier transform infrared spectroscopy (FT-IR) spectra from 400 to 4000 cm⁻¹ (NICOLET iN10 MX, Thermo Scientific, USA). The morphology of the composites were observed using scanning electron microscope (SEM, Hitachi S-4800, Japan) with 10 kV accelerating voltage, and the Brunauer-Emmett-Teller (BET) specific surface areas (S_BET) were measured with ASAP 2020M + C (Micromeritics, USA) from N₂ adsorption method.

Procedures and analysis

In a typical sacrificial batch experiment, a 20 mL cylindrical glass vessel was fully filled with 0.10 mmol L⁻¹ TCE, appropriate amounts of H₂O₂ and activators (i.e., nZVI, BC or nZVI/BC), and the vessel was tightly sealed with Teflon reaction head successively. Then, the reaction was initiated in a rotary shaker at 298 K and 150 rpm. Control test was also carried out under the same condition without H₂O₂ and activators. Samples were taken at the desired reaction time intervals and filtered through 0.2 μm membrane prior to the analysis. Samples were conducted in triplicate and the mean value was obtained.

The concentration of TCE was analyzed by headspace Gas Chromatograph Mass Spectrometer (GC-MS, Agilent 7890A and 5975C) using DB-624 chromatographic column (30.0 m × 0.25 mm × 1.4 μm). Dissolved ferrous ion was quantified through 1,10-Phenanthrolinemonohydrate Spectrophotometry by using a Cary 50 UV-vis spectrophotometer (Varian Cary 50, USA). The concentration of H₂O₂ was measured with the DPD method using UV-vis spectrophotometer (Varian Cary 50, USA). The generated radical species was detected with electron spin resonance spectrometer (Bruker ESR 300E, Germany) with microwave bridge (receiver gain, 1 × 10⁵; modulation amplitude, 2 Gauss; microwave power, 10 mW; modulation frequency, 100 kHz) using DMPO as radical spin-trapping reagent, and fluorescence spectra were measured on fluorescence spectrophotometer (Jasco FP-6200, Japan). Total organic carbon (TOC) was recorded with a multi N/C model TOC analyzer (Analytik Jena, multi N/C 2100, Germany).