Degradation of 2,4-Dinitrophenol in Aqueous Solution by Microscale Fe 0 /H 2 O 2 /O 3 Process

Abstract

In this study, a microscale Fe⁰ (mFe⁰)/H₂O₂/O₃ process was set up to degrade 2,4-dinitrophenol (DNP) in aqueous solution. First, key operating parameters including initial solution pH, mFe⁰ dosage, O₃ flow rate, and H₂O₂ dosage on Chemical Oxygen Demand (COD) removal efficiency were optimized through single-factor experiments. Under optimal conditions, COD removal efficiency reached 80.2%, and Biochemical Oxygen Demand (BOD₅)/COD (B/C) ratio was elevated from 0 to 0.61 after 15 min treatment. Moreover, compared with control experiments (i.e., mFe⁰, O₃, H₂O₂, mFe⁰/O₃, mFe⁰/H₂O₂, O₃/H₂O₂, mFe⁰/H₂O₂/O₂), the mFe⁰/H₂O₂/O₃ process exerted better performance for degradation of DNP in aqueous solution due to the strong synergistic effect between mFe⁰, H₂O₂, and O₃. In addition, degradation and transformation of DNP were also analyzed. Finally, a reaction mechanism of mFe⁰/H₂O₂/O₃ process was proposed. It can be concluded that efficient mFe⁰/H₂O₂/O₃ process mainly resulted from the combination effects of heterogeneous and homogeneous catalytic ozonation process, Fenton and Fenton-like oxidation process, and direct oxidation process by O₃. Therefore, the mFe⁰/H₂O₂/O₃ process could be proposed as a high-efficient treatment technology for removal of toxic refractory DNP from wastewater.

Introduction

Alkyl dinitrophenols are widely acknowledged as a group of toxic refractory contaminants, which show low biodegradability and pose serious risks to human health and environment (Shiraishi et al., 2015). 2,4-Dinitrophenol (DNP) is a typical chemical compound of this class of toxic compounds, and it has been extensively used in petrochemical industry as polymerization inhibitor for vinyl aromatics and in agriculture as a pesticide (Wang et al., 2009; Ma et al., 2014). Although many benefits have been obtained by the application of DNP, it also has some undesirable side effects. The carcinogenic nature of DNP creates significant health risk, and it can restrict the cell growth even at much lower concentration levels (1 ppm) (Cao and Shiraishi, 2010).

DNP has been identified as priority pollution agent by the United States Environmental Protection Agency (USEPA), and the recommend restriction on the concentration in natural water is below 10 ng/L (Pillai and Gupta, 2015). Due to its low-molecular weight and high toxicity, it is resistant to biological degradation and can be hardly destroyed by conventional technologies for wastewater treatment (Wang et al., 2010a). If wastewaters containing DNP were released into environment, it would pose a severe threat to humans and wildlife.

Over the past decade, the advanced oxidation processes (AOPs), such as Fenton's reagent (Wei et al., 2013), electro-Fenton process (Zhu et al., 2011), photocatalytic oxidation (Zhang et al., 2017a), ozonation process (Nawrocki and Kasprzyk-Hordern, 2010; Zhang et al., 2017b), and zero-valent iron (ZVI, Fe⁰)/air process (Xiong et al., 2015), have been commonly recognized as the potential alternative techniques for the degradation of toxic and/or refractory pollutants in water and wastewater based on the generation of hydroxyl radical (HO^•, E⁰ = 2.8 V).

Many researches about the treatment processes of DNP have been reported in the literatures, such as photocatalytic oxidation (Shiraishi et al., 2015; Zhang et al., 2016; Cai et al., 2017; Chen et al., 2017), photo-assisted Fenton oxidation (Wang et al., 2010a), and electrochemical oxidation (Pillai and Gupta, 2015). However, the application of these processes is still limited due to the operating cost or treatment efficiency. Hence, it is necessary to develop a more cost-effective, robust, and feasible treatment method to degrade DNP before it was discharged into the environment.

As an inexpensive, widely available, and environmental friendly reactive metal, Fe⁰ has typically been used for the treatment of toxic and refractory pollutants contained in groundwater or wastewater (Lee et al., 2013; Liu et al., 2013; Gong et al., 2015; Zhang et al., 2017c). The degradation/removal of pollutants by Fe⁰ is mainly attributed to the common effect of oxidation, reduction, coprecipitation, adsorption, and others (Yoon et al., 2011; Bautitz et al., 2012). Recently, there has been a growing interest in Fe⁰-mediated AOPs using microscale Fe⁰ (mFe⁰) particles for the treatment of toxic and refractory pollutants; mFe⁰/air, mFe⁰/O₃, mFe⁰/H₂O₂, and mFe⁰/persulfate (PS) were developed and extensively reported in the literatures (Hussain et al., 2012; Martins et al., 2012; Xiong et al., 2015; Xiong et al., 2016a).

These are physicochemical processes based on the generation of HO^• or sulfate free radical (SO₄^•−, E⁰ = 2.6 V), which has strong oxidation capacity and could degrade refractory organic pollutants efficiently. Meanwhile, ozone-based AOPs such as Fe⁰/O₃, Fe²⁺/O₃, O₃/H₂O₂, and O₃/UV also have been widely investigated and proven to be an efficient method for toxic and refractory wastewater treatment (Zeng et al., 2012; Qi et al., 2016; Xiong et al., 2016a; Srithep and Phattarapattamawong, 2017; Ji et al., 2018).

In our previous studies, we also found synergistic effects between mFe⁰ and O₃ in removal of refractory pollutants (Xiong et al., 2016b). In mFe⁰/O₃ system, ozone can be directly reduced by Fe⁰ and the generated iron ions could catalyze the ozone decomposition and generate more radicals. In particular, the organic acids generated in the decomposition of pollutants by direct ozonation or catalytic ozonation process also could accelerate the iron corrosion, which could facilitate the catalytic ozonation process in return (Nawrocki and Kasprzyk-Hordern, 2010; Wang and Bai, 2017). The stoichiometric reactions for catalytic ozonation by Fe⁰ and its corrosion products are expressed in Equations (1 )–(6) (Beltrán et al., 2005; Xiong et al., 2016a; Xiong et al., 2016b; Wang and Bai, 2017). \documentclass{aastex}\usepackage{amsbsy}\usepackage{amsfonts}\usepackage{amssymb}\usepackage{bm}\usepackage{mathrsfs}\usepackage{pifont}\usepackage{stmaryrd}\usepackage{textcomp}\usepackage{portland, xspace}\usepackage{amsmath, amsxtra}\usepackage{upgreek}\pagestyle{empty}\DeclareMathSizes{10}{9}{7}{6}\begin{document} \begin{align*} { \rm{Fe + 2}}{{ \rm{O}}_{ \rm{3}}} \to { \rm{F}}{{ \rm{e}}^{{ \rm{2 + }}}}{ \rm{ + 2O}}_{ \rm{3}}^{ \bullet - } \tag{1} \end{align*} \end{document} \documentclass{aastex}\usepackage{amsbsy}\usepackage{amsfonts}\usepackage{amssymb}\usepackage{bm}\usepackage{mathrsfs}\usepackage{pifont}\usepackage{stmaryrd}\usepackage{textcomp}\usepackage{portland, xspace}\usepackage{amsmath, amsxtra}\usepackage{upgreek}\pagestyle{empty}\DeclareMathSizes{10}{9}{7}{6}\begin{document} \begin{align*} { \rm{F}}{{ \rm{e}}^{{ \rm{2 + }}}}{ \rm{ + }}{{ \rm{O}}_{ \rm{3}}} \to { \rm{F}}{{ \rm{e}}^{{ \rm{3 + }}}}{ \rm{ + O}}_{ \rm{3}}^{ \bullet - } \tag{2} \end{align*} \end{document} \documentclass{aastex}\usepackage{amsbsy}\usepackage{amsfonts}\usepackage{amssymb}\usepackage{bm}\usepackage{mathrsfs}\usepackage{pifont}\usepackage{stmaryrd}\usepackage{textcomp}\usepackage{portland, xspace}\usepackage{amsmath, amsxtra}\usepackage{upgreek}\pagestyle{empty}\DeclareMathSizes{10}{9}{7}{6}\begin{document} \begin{align*} {{ \rm{H}}^{ \rm{ + }}}{ \rm{ + O}}_3^{ \bullet - } \to { \rm{H}}{{ \rm{O}}_{ \rm{3}}} \to { \rm{HO}}_{}^ \bullet + {{ \rm{O}}_{ \rm{2}}} \tag{3} \end{align*} \end{document} \documentclass{aastex}\usepackage{amsbsy}\usepackage{amsfonts}\usepackage{amssymb}\usepackage{bm}\usepackage{mathrsfs}\usepackage{pifont}\usepackage{stmaryrd}\usepackage{textcomp}\usepackage{portland, xspace}\usepackage{amsmath, amsxtra}\usepackage{upgreek}\pagestyle{empty}\DeclareMathSizes{10}{9}{7}{6}\begin{document} \begin{align*} { \rm{F}}{{ \rm{e}}^{{ \rm{2 + }}}}{ \rm{ + O}}_3^{} \to { \rm{Fe}}{{ \rm{O}}^{{ \rm{2 + }}}} + {{ \rm{O}}_2} \tag{4} \end{align*} \end{document} \documentclass{aastex}\usepackage{amsbsy}\usepackage{amsfonts}\usepackage{amssymb}\usepackage{bm}\usepackage{mathrsfs}\usepackage{pifont}\usepackage{stmaryrd}\usepackage{textcomp}\usepackage{portland, xspace}\usepackage{amsmath, amsxtra}\usepackage{upgreek}\pagestyle{empty}\DeclareMathSizes{10}{9}{7}{6}\begin{document} \begin{align*} { \rm{Fe}}{{ \rm{O}}^{{ \rm{2 + }}}}{ \rm{ + }}{{ \rm{H}}_2}{ \rm{O}} \to { \rm{F}}{{ \rm{e}}^{{ \rm{3 + }}}} + { \rm{H}}{{ \rm{O}}^ \bullet } + { \rm{O}}{{ \rm{H}}^{ \rm{ - }}} \tag{5} \end{align*} \end{document} \documentclass{aastex}\usepackage{amsbsy}\usepackage{amsfonts}\usepackage{amssymb}\usepackage{bm}\usepackage{mathrsfs}\usepackage{pifont}\usepackage{stmaryrd}\usepackage{textcomp}\usepackage{portland, xspace}\usepackage{amsmath, amsxtra}\usepackage{upgreek}\pagestyle{empty}\DeclareMathSizes{10}{9}{7}{6}\begin{document} \begin{align*} { \rm{F}}{{ \rm{e}}^{{ \rm{3 + }}}}{ \rm{ + }}{{ \rm{O}}_{ \rm{3}}} \to { \rm{Fe}}{{ \rm{O}}^{{ \rm{2 + }}}}{ \rm{ + HO}}_{}^ \bullet + {{ \rm{O}}_{ \rm{2}}}{ \rm{ + }}{{ \rm{H}}^{ \rm{ + }}} \tag{6} \end{align*} \end{document}

On the basis of Fe⁰-based AOPs (e.g., Fe⁰/H₂O₂, Fe⁰/O₃) and ozone-based AOPs (e.g., Fe⁰/O₃, O₃/H₂O₂), in this study, an mFe⁰/H₂O₂/O₃ process was set up to further improve the treatment efficiency of DNP through their synergistic effects. First, effects of key operating parameters including initial solution pH value, mFe⁰ dosage, O₃ flow rate, and H₂O₂ dosage on the Chemical Oxygen Demand (COD) removal efficiency of DNP in aqueous solution were optimized through single-factor experiments. Then, the control experiments were set up to investigate the synergistic effects and the superiority of mFe⁰/H₂O₂/O₃ process. Moreover, the degradation and transformation of DNP were analyzed according to the present literatures and analytic data of COD, B/C ratio, UV-Vis spectra, and transformation of nitrogen. Finally, a possible reaction mechanism of the mFe⁰/H₂O₂/O₃ process was discussed thoroughly. The aim of this work was to assess the feasibility of mFe⁰/H₂O₂/O₃ process on the degradation of DNP, which is a kind of typical toxic and refractory pollutant.

Materials and Methods

Chemicals

DNP (99%), hydrogen peroxide (H₂O₂, 30% w/v), mFe⁰, sodium hydroxide, and sulfuric acid were purchased from Chengdu Kelong Chemical Reagent Factory. All chemicals used in the experiment were of analytical grade and used as received without further purification. Deionized water was used throughout the whole experiment process.

Experimental procedure

DNP aqueous solution (500 mg/L) was prepared by simple dissolution in deionized water. Its pH, COD, Biochemical Oxygen Demand (BOD₅), and BOD₅/COD (B/C) ratio were 3.3 ± 0.1 mg/L, 595 ± 5 mg/L, 0 mg/L, and 0, respectively. These characteristics suggest that it is hard to be treated directly through the conventional processes. Batch degradation experiments of DNP were carried out in a 500-mL flat bottom beaker. Ozone was produced onsite from pure oxygen through an ozone generator, and the inlet concentration of ozone was about 46 mg/L. In each batch experiment, 300 mL DNP was placed in a 500-mL beaker, meanwhile the desired mFe⁰ and H₂O₂ dosages were added into the beaker, and the reaction solution was stirred by a mechanical stirrer with a speed of 300 rpm. Simultaneously, O₃ with a desired flow rate was continuously dispersed into the reaction solution through a gas diffuser.

To confirm the superiority of mFe⁰/H₂O₂/O₃ process and the synergistic effect between mFe⁰, H₂O₂, and O₃, seven control experiments were set up: (a) mFe⁰, (b) H₂O₂, (c) O₃, (d) mFe⁰/O₃, (e) mFe⁰/H₂O₂, (f) O₃/H₂O₂, and (g) mFe⁰/H₂O₂/O₂. And other operating conditions of control experiments were in accordance with the optimal conditions of the mFe⁰/H₂O₂/O₃ process. In particular, mFe⁰ dosage of (a) mFe⁰, (d) mFe⁰/O₃, and (e) mFe⁰/H₂O₂ was same to that of the mFe⁰/H₂O₂/O₃ system. H₂O₂ dosage of (c) H₂O₂, (e) mFe⁰/H₂O₂, and (f) O₃/H₂O₂ was same to that of the mFe⁰/H₂O₂/O₃ system. O₃ and O₂ flow rate of (c) O₃, (d) mFe⁰/O₃, (f) O₃/H₂O₂, and (g) mFe⁰/H₂O₂/O₂ was same to that of the mFe⁰/H₂O₂/O₃ system.

The samples were withdrawn at predetermined time intervals and diluted by deionized water and filtered through a poly tetra fluoroethylene (PTFE) syringe filter disc (0.45 μm). And then, the COD, BOD₅, total iron ion concentration, and H₂O₂ utilization of the samples were measured.

The whole experiment process was performed at a running temperature of 25 ± 1°C by heating in water bath. Water samples were withdrawn at regular intervals and filtered by using a 0.45-μm one-off syringe filter. Solution pH was precisely adjusted by H₂SO₄ and NaOH in all reactions for the comparison. All the experiments were conducted in triplicate, and the error bar represented the standard deviation of the replicate experimental data.

Analytical method

Surface morphology and elementary composition of fresh and reacted mFe⁰ particles and flocculation formed by the mFe⁰/H₂O₂/O₃ process were observed by scanning electron microscopy (SEM, JSM-7500F; JEOL Ltd., Japan) and energy-dispersive spectrometry (EDS; Oxford Instruments). On the basis of the elemental composition analysis, its compound composition was further investigated by X-ray diffraction (XRD; PANalytic B.V., Holland). X-ray photoelectron spectroscopy (XPS, AXIS Ultra DLD; Kratos Co., United Kingdom) was used to identify the chemical states of Fe and O on the surface of reacted mFe⁰ of the mFe⁰/H₂O₂/O₃ process. The UV-Vis absorption spectra of the effluent samples were performed in 10-mm quartz cuvettes by UV-Vis spectrophotometer (Shimazu, Japan), and the UV-Vis spectra were recorded from 190 to 500 nm using distilled water as blank.

The total iron ion leaching was detected with phenanthroline spectrophotometry (UV1700; Shimazu, Japan). The iodometric method was used to monitor the gas ozone concentration (Roth and Sullivan, 2002). The concentration of H₂O₂ was analyzed using potassium titanium(IV) oxalate spectrophotometric method (Sellers, 1980). The generated NO₂⁻ and NO₃⁻ were analyzed using an ion chromatography (IC, ICS1100; Dionex). BOD₅, COD, and pH of the samples were measured by BOD₅ analyzer (OxiTop IS12; WTW, Germany), COD rapid analyzer (Lianhua, China), and PHS-3C meter (Rex, China), respectively.

Results and Discussion

Optimization of operation parameters

To obtain optimal conditions, effects of key operating parameters including initial solution pH value, mFe⁰ dosage, O₃ flow rate, and H₂O₂ dosage on the COD removal efficiency of DNP in aqueous solution were optimized through single-factor experiment. In addition, variations of the effluent pH also were monitored carefully during the experiment process.

Effect of initial solution pH

It is well known that solution pH value has significant influence on catalytic ozonation and Fenton and Fenton-like oxidation process (Remya and Lin, 2011; Martins et al., 2012; Pan et al., 2012; Wang and Bai, 2017). Therefore, it is quite important to investigate the effect of initial solution pH on the COD removal efficiency of DNP aqueous solution by the mFe⁰/H₂O₂/O₃ process. As shown in Fig. 1a, the COD removal efficiency remarkably enhanced from 14.9% to 57.2% when the initial solution pH deceased from 10.0 to 7.0, and then, it gradually increased to 76.4% when the initial solution pH decreased to 2.0.

FIG. 1.

Effect of (a) initial pH, (b) mFe⁰ dosage, (c) O₃ flow rate, and (d) H₂O₂ dosage on the COD removal efficiencies and effluent pH after mFe⁰/H₂O₂/O₃ treatment. COD, Chemical Oxygen Demand.

The solution pH is an important parameter for mFe⁰ process because the lower solution pH can significantly accelerate iron corrosion rate and generate more corrosion products (e.g., Fe²⁺, Fe³⁺). These corrosion products can catalyze ozone decomposition [Eqs. (2 )–(6)] and take part in the Fenton reaction [Eq. (8)] to generate enough HO^•, consequently result in higher COD removal efficiency of DNP aqueous solution in the mFe⁰/H₂O₂/O₃ process. In addition, the lower solution pH can also facilitate the formation of Fenton-like reaction under oxic condition [Eqs. (7) and (8)], which further improve the COD removal efficiency. With the increase of initial solution pH, Fenton reaction and catalytic ozonation processes by Fe²⁺/Fe³⁺ were gradually inhibited due to the decrease of iron corrosion rate.

Fenton-like reaction was also inhibited due to the decrease of H⁺ concentration. Therefore, a satisfied COD removal efficiency cannot be obtained in neutral and alkaline conditions. It is well known that the increase of solution pH can accelerate ozone decomposition to generate HO^• [Eq. (9)] (Zhao et al., 2017). However, a converse variation trend appeared because HO^• produced by ozone decomposition did not play a leading role in the mFe⁰/H₂O₂/O₃ process. Therefore, the higher COD removal efficiency can be attained in lower initial solution pH. However, when the initial solution pH was below 2.0, the COD removal efficiency has a significant drop (Fig. 1a).

The COD removal efficiency at extremely low solution pH is undesirable, which can be explained from the following four aspects: (i) the excess H⁺ ions can act as HO^• radical scavenger (Zeng et al., 2012); (ii) the lower pH could yield the excess Fe²⁺ that also can act as HO^• radical scavenger (Song et al., 2016); (iii) the excess H⁺ ions would react with H₂O₂ to form H₃O₂⁺, leading to an improved stability of H₂O₂ and weaken the COD removal efficiency (Bokare and Choi, 2014); and (iv) the excess H⁺ ions could restrain the decomposition of ozone (Kasprzyk-Hordern et al., 2003). As a result, the optimal initial solution pH of 2.5 was selected to optimize the following parameters.

In addition, Fig. 1a also shows the effect of initial solution pH on the effluent pH. In particular, the effluent pH was higher than the initial solution pH when the initial solution pH was below 3.3. On the contrary, the effluent pH was lower than the initial solution pH when the initial solution pH was above 3.3. The results can be explained from the following two aspects: (i) the catalytic ozonation process and Fenton and Fenton-like reaction could be limited seriously under the neutral or alkaline condition; thus, the sole ozonation and O₃/H₂O₂ process might play a leading role in the degradation of pollutants. The acidic intermediates (e.g., carboxylic acids) might be generated during the sole ozonation and O₃/H₂O₂ process due to the inefficient oxidation, which decreased the pH of reaction solution. (ii) Under acidic condition, H⁺ ions could be rapidly consumed by mFe⁰, consequently result in the increased of effluent pH (Son et al., 2009). \documentclass{aastex}\usepackage{amsbsy}\usepackage{amsfonts}\usepackage{amssymb}\usepackage{bm}\usepackage{mathrsfs}\usepackage{pifont}\usepackage{stmaryrd}\usepackage{textcomp}\usepackage{portland, xspace}\usepackage{amsmath, amsxtra}\usepackage{upgreek}\pagestyle{empty}\DeclareMathSizes{10}{9}{7}{6}\begin{document} \begin{align*} { \rm{F}}{{ \rm{e}}^{ \rm{0}}}{ \rm{ + O}}_2^{} + 2{{ \rm{H}}^ + } \to { \rm{F}}{{ \rm{e}}^{{ \rm{2 + }}}} + {{ \rm{H}}_2}{{ \rm{O}}_2} \tag{7} \end{align*} \end{document} \documentclass{aastex}\usepackage{amsbsy}\usepackage{amsfonts}\usepackage{amssymb}\usepackage{bm}\usepackage{mathrsfs}\usepackage{pifont}\usepackage{stmaryrd}\usepackage{textcomp}\usepackage{portland, xspace}\usepackage{amsmath, amsxtra}\usepackage{upgreek}\pagestyle{empty}\DeclareMathSizes{10}{9}{7}{6}\begin{document} \begin{align*} { \rm{F}}{{ \rm{e}}^{2 + }}{ \rm{ + }}{{ \rm{H}}_{ \rm{2}}}{ \rm{O}}_2^{} \to { \rm{F}}{{ \rm{e}}^{{ \rm{3 + }}}}{ \rm{ + O}}{{ \rm{H}}^ - } + { \rm{HO}}_{}^ \bullet \tag{8} \end{align*} \end{document} \documentclass{aastex}\usepackage{amsbsy}\usepackage{amsfonts}\usepackage{amssymb}\usepackage{bm}\usepackage{mathrsfs}\usepackage{pifont}\usepackage{stmaryrd}\usepackage{textcomp}\usepackage{portland, xspace}\usepackage{amsmath, amsxtra}\usepackage{upgreek}\pagestyle{empty}\DeclareMathSizes{10}{9}{7}{6}\begin{document} \begin{align*} {{ \rm{O}}_3}{ \rm{ + OH}}_{}^ - \to { \rm{HO}}_{ \rm{2}}^ \bullet { + ^ \bullet }{ \rm{O}}_2^ - \to { \rm{HO}}_{}^ \bullet \tag{9} \end{align*} \end{document} \documentclass{aastex}\usepackage{amsbsy}\usepackage{amsfonts}\usepackage{amssymb}\usepackage{bm}\usepackage{mathrsfs}\usepackage{pifont}\usepackage{stmaryrd}\usepackage{textcomp}\usepackage{portland, xspace}\usepackage{amsmath, amsxtra}\usepackage{upgreek}\pagestyle{empty}\DeclareMathSizes{10}{9}{7}{6}\begin{document} \begin{align*} { \rm{F}}{{ \rm{e}}^{ \rm{0}}}{ \rm{ + }}2{ \rm{F}}{{ \rm{e}}^{{ \rm{3 + }}}} \to 3{ \rm{F}}{{ \rm{e}}^{{ \rm{2 + }}}} \tag{10} \end{align*} \end{document}

Effect of mFe⁰ dosage

Figure 1b shows the effect of mFe⁰ dosage on COD removal efficiency by the mFe⁰/H₂O₂/O₃ process. It is apparent that the lower COD removal efficiency (47.9%) was obtained after 15 min treatment by the O₃/H₂O₂ process without mFe⁰, whereas it was improved from 47.9% to 72.6% with the increase of mFe⁰ dosage from 0 to 3 g/L, and then, it would not be further significantly enhanced when the mFe⁰ dosage was above 3 g/L.

The results can be explained from the following two aspects: (i) the increase of mFe⁰ dosage could facilitate the formation of iron corrosion products; a large amount of mFe⁰ and its corrosion products can enhance the utilization rate of O₃ and catalyze the decomposition of O₃ to generate enough stronger oxidants (i.e., HO^• and O^•) that can easily decompose pollutants in reaction solution (Sui et al., 2010; Hussain et al., 2012; Chen et al., 2016; Oh et al., 2016). In addition, mFe⁰ and its corrosion products can also take part in Fenton or Fenton-like reaction and produce more HO^•, resulting in the higher COD removal efficiency. (ii) Further increase of mFe⁰ dosage had no significant contribution to COD removal efficiency because the activators (i.e., H₂O₂ and O₃) might become the limited factor when the excessive mFe⁰ was added to reaction solution. Therefore, the optimal mFe⁰ dosage of 3.0 g/L was selected to carry out the following experiments.

Figure 1a also shows the effect of mFe⁰ dosage on the effluent pH. The effluent pH of the mFe⁰/H₂O₂/O₃ process gradually increased to 4.6 with the increase of mFe⁰ dosage from 0 to 5 g/L, and then, it would not be further significantly enhanced when the mFe⁰ dosage was above 5 g/L. The results indicate that the H⁺ ions could be rapidly consumed by mFe⁰ when a few of mFe⁰ powders were added in this treatment system. With the increase of mFe⁰ dosage, the accumulated acids and H⁺ could be consumed gradually and generate a plenty iron corrosion products, which could increase the effluent pH.

Effect of O₃ flow rate

Effect of O₃ flow rate on the COD removal efficiency by the mFe⁰/H₂O₂/O₃ process is shown in Fig. 1c. It is apparent that the mFe⁰/H₂O₂ process without O₃ only led to 22.8% of COD removal efficiency. It suggests that the mFe⁰/H₂O₂ process is a relatively inefficient process for pollutants degradation. It could also be observed from Fig. 1c that COD removal efficiency increased rapidly from 22.8% to 72.6% when O₃ flow rate elevated from 0 to 0.3 L/min after 15 min treatment by the mFe⁰/H₂O₂/O₃ process. And then, the COD removal did not further significantly improve when the O₃ flow rate was further increased to 0.6 L/min.

The results can be explained from the following three aspects: (i) the increase of O₃ flow rate would improve the amount of dissolved ozone in reaction solution when O₃ flow rate was below 0.3 L/min, which facilitate the catalytic ozonation process. More HO^• would be generated when more O₃ was introduced into the reaction solution, resulting in higher COD removal efficiency. (ii) The generated gas contained O₃ and O₂ because ozone was produced onsite from dry O₂ through an ozone generator. Increasing the aeration rate also would increase the amount of O₂ in reaction solution, which would facilitate Fenton-like reaction to further improve the COD removal efficiency. (iii) The further increase of ozone flow rate (>0.3 L/min) had no significant contribution to COD removal because the O₃ and O₂ concentrations in the liquid phase had approached to saturation at a fixed temperature (Pan et al., 2012). Thus, the ozone flow rate of 0.3 L/min was selected as the optimal value in the follow-up study.

In addition, Fig. 1c also shows that the effluent pH of the mFe⁰/H₂O₂/O₃ process decreased gradually from 4.7 to 3.6 when O₃ flow rate increased from 0 to 0.4 L/min, and then, it did not further significantly decrease when the O₃ flow rate was above 0.4 L/min. With the increase of O₃ flow rate, the mFe⁰/O₃/H₂O₂ process might accumulate more organic acids, which led to the decrease of effluent pH.

Effect of H₂O₂ dosage

Figure 1d shows the effect of H₂O₂ dosage on COD removal efficiency by the mFe⁰/H₂O₂/O₃ process. The COD removal efficiency rapidly increased from 55.2% to 80.2% with an increase of H₂O₂ dosage from 0 to 25 mmol/L. And then, it did not further increase when H₂O₂ dosage was above 25 mmol/L. The results indicate that COD removal efficiency was significantly influenced by H₂O₂ dosage.

It can be explained from the following two aspects: (i) higher H₂O₂ dosage in reaction solution was beneficial for the Fenton reaction and O₃/H₂O₂ process to generate HO^• that accounted for the obvious enhancement of COD removal efficiency (Safarzadeh-Amiri, 2001; Segura et al., 2012; Medellin-Castillo et al., 2013; Fukuchi et al., 2014). (ii) The further increase of H₂O₂ dosage had no significant contribution to COD removal because other activators might become the limited factor when the excessive H₂O₂ was added in reaction solution. Therefore, 40 mmol/L of H₂O₂ dosage was selected as the optimal H₂O₂ dosage.

In addition, Fig. 1d also shows that the effluent pH (3.1–4.0) of the mFe⁰/H₂O₂/O₃ process was not affected significantly by the different H₂O₂ dosage (0–40 mmol/L). Their effluent pH only decreased a little (4.0–3.1) with the increase of H₂O₂ dosage from 0 to 40 mmol/L.

Finally, the optimal operating parameters including initial pH (2.5), mFe⁰ dosage (3.0 g/L), O₃ flow rate (0.3 L/min), and H₂O₂ dosage (25 mmol/L) were obtained according to the above semibatch experiments.

Control experiments

To evaluate the advantage of mFe⁰/O₃/H₂O₂ process and the synergetic effect between mFe⁰, H₂O₂, and O₃, seven control experiments including (a) mFe⁰, (b) H₂O₂, (c) O₃, (d) mFe⁰/O₃, (e) mFe⁰/H₂O₂, (f) O₃/H₂O₂, and (g) mFe⁰/H₂O₂/O₂ processes were carried out under the same conditions (i.e., initial pH of 2.5, mFe⁰ dosage of 3.0 g/L, O₃ flow rate of 0.3 L/min, O₂ flow rate of 0.3 L/min, H₂O₂ dosage of 25 mmol/L, and treatment time of 15 min). According to Equations (3), (4), (6), (13), and (15), O₂ would be generated during the catalytic ozonation process. Meanwhile, the gas dispersed into the reaction solution contained O₃ and O₂. Therefore, the mFe⁰/H₂O₂/O₂ process was used to evaluate the contribution of O₂ in the mFe⁰/H₂O₂/O₃ process.

The performance of mFe⁰/H₂O₂/O₃ process and control experiments were investigated comparatively through analyzing the COD removal efficiency and total iron ion concentration (i.e., Fe²⁺, Fe³⁺, and flocculent precipitate) in different reaction times, biodegradability improvement, and H₂O₂ utilization.

COD removal during the treatment process

Variation of COD removal efficiency with reaction time in the eight different experiments was shown in Fig. 2a. It is clear that all the COD removal efficiencies obtained by the mFe⁰/H₂O₂/O₃ process were much higher than those of seven control experiments during 15 min treatment. The sequence of COD removal efficiency after 15 min treatment process in the different systems follow the trend that mFe⁰/H₂O₂/O₃ (80.2%) > mFe⁰/O₃ (54.4%) > mFe⁰/H₂O₂/O₂ (52.6%) > mFe⁰/H₂O₂ (44.6%) > O₃/H₂O₂ (36.8%) > sole O₃ (23.3%) > mFe⁰ (9.0%) > H₂O₂ (0%). It can be concluded that DNP was rarely oxidized by H₂O₂ due to the limited oxidation capability of H₂O₂ (E^θ = 1.78 V).

FIG. 2.

COD removal efficiency (a), B/C ratio (b), total iron ion concentration (c), and H₂O₂ utilization (d) obtained by the different systems ([DNP] = 500 mg/L, mFe⁰ dosage = 3 g/L, H₂O₂ dosage = 25 mmol/L, O₃ flow rate = 0.3 L/min, O₂ flow rate = 0.3 L/min, initial solution pH = 2.5). DNP, 2,4-dinitrophenol.

The COD removal efficiency of the mFe⁰/H₂O₂/O₂ process was higher than that of the mFe⁰/H₂O₂ process, which indicates that O₂ has contribution to COD removal in the mFe⁰/H₂O₂/O₃ process due to the formation of Fenton-like reaction [Eqs. (7) and (8)]. In particular, the sum (32.3%) of COD removal efficiencies of processes by mFe⁰, H₂O₂, and O₃ alone was much lower than that (80.2%) obtained by the mFe⁰/H₂O₂/O₃ process. Furthermore, the sum (54.4%, 66.9%, and 44.8%, respectively) of COD removal efficiencies of mFe⁰/O₃ and sole H₂O₂ process, mFe⁰/H₂O₂ and sole O₃ process, O₃/H₂O₂ and sole mFe⁰ process was also lower than that (80.2%) obtained by the mFe⁰/H₂O₂/O₃ process. The results can confirm that there is extremely strong synergetic effect between mFe⁰, H₂O₂, and O₃ in the degradation of DNP.

Biodegradability

Figure 2b shows the B/C ratios of the raw DNP solution, effluent of mFe⁰/H₂O₂/O₃ process, and seven different control experiments. Ordinarily, the wastewater is considered to be biodegradable when its B/C ratio is higher than 0.30. With a B/C ratio of 0, the DNP wastewater was clearly biorefractory. After the 15 min treatment process, the sequence of B/C ratios obtained by the different processes are as follows: mFe⁰/H₂O₂/O₃ (0.61) > mFe⁰/H₂O₂/O₂ (0.25) > mFe⁰/H₂O₂ (0.09) > mFe⁰/O₃ (0.06) > O₃ (0.04) > O₃/H₂O₂ (0.03) > mFe⁰ (0.02) > H₂O₂ (0).

The B/C ratio of the effluent of mFe⁰/H₂O₂/O₃ process was much higher than that of seven control experiments. In particular, only the B/C ratio of the effluent of mFe⁰/H₂O₂/O₃ process was higher than 0.30. The results indicate that DNP could be effectively degraded or transformed by mFe⁰/H₂O₂/O₃ process, and its biodegradability was improved significantly. It also further indicates that there was a strong synergistic reaction in the mFe⁰/H₂O₂/O₃ process.

Total iron ion concentration

Figure 2c shows the variation of total iron ions (i.e., Fe²⁺, Fe³⁺, and flocculent precipitate) concentration in reaction solution during 15 min treatment process by the different treatment processes. It is clear that O₃, O₂, and H₂O₂ have considerable abilities to promote iron corrosion and facilitate the releasing of iron ions during the treatment of DNP. In addition, the total iron ions concentrations of mFe⁰/H₂O₂/O₃ process were much higher than those of mFe⁰/H₂O₂/O₂ process, mFe⁰/H₂O₂ process, mFe⁰/O₃ process, and mFe⁰ process during 15 min treatment. In particular, it can be observed that the total iron ions concentration in the mFe⁰/H₂O₂/O₃ process rapidly increased to about 475.6 mg/L after 15 min treatment, which was much higher than that of the mFe⁰/H₂O₂/O₂ process (about 347.9 mg/L), mFe⁰/H₂O₂ process (about 287.1 mg/L), mFe⁰/O₃ process (about 317.6 mg/L), and mFe⁰ process (about 199.3 mg/L).

The corrosion products could facilitate the generation of HO^• by catalytic ozonation and formation of the Fenton reaction. The results suggest that the corrosion of mFe⁰ could be improved remarkably by the combination O₃, O₂, and H₂O₂, which results in higher reaction reactivity of mFe⁰/H₂O₂/O₃ process. In addition, because of the variation of solution pH was lower than 3.5 during the 15 min treatment of mFe⁰/H₂O₂/O₃ process, the generation of flocculent precipitate after treatment by the mFe⁰/H₂O₂/O₃ process was very few. Furthermore, it will not cause secondary pollution in the environment and could be dealt with promptly and safely by landfill disposal, iron making, and steelmaking.

The utilization of H₂O₂

Figure 2d shows the H₂O₂ utilization by different treatment processes after 15 min treatment. The sequence of H₂O₂ utilization in the different systems follow the trend that mFe⁰/H₂O₂/O₃ (96.5%) > mFe⁰/H₂O₂/O₂ (88.4%) > mFe⁰/H₂O₂ (85.6%) > O₃/H₂O₂ (15.6%) > H₂O₂ (0). It is clear that the H₂O₂ conversion efficiency by the mFe⁰/H₂O₂/O₃ process after 15 min treatment was higher than those of the mFe⁰/H₂O₂ process, mFe⁰/H₂O₂/O₂ process, O₃/H₂O₂ process, and sole H₂O₂ process. In addition, H₂O₂ can also be formed through Equation (7). Therefore, the H₂O₂ consumption of mFe⁰/H₂O₂/O₃ and mFe⁰/H₂O₂/O₂ processes includes the added and generated H₂O₂. It further indicates that H₂O₂ can have higher conversion efficiency in the mFe⁰/H₂O₂/O₃ process.

UV-Vis spectral analysis and transformation of nitrogen

Figure 3a shows the real-time UV-Vis absorption spectra changes of DNP in aqueous solution during the mFe⁰/H₂O₂/O₃ process under the optimal conditions. The UV-Vis absorbance peak of influent was detected from 190 to 500 nm, which shows that the UV characteristic absorbance peak of DNP was mainly at 212 nm and 260 nm. In particular, the absorption peaks at 212 and 260 nm disappeared only after 1 min treatment. And then, the absorbance decreased progressively following the 14 min treatment. The results indicate that the mFe⁰/H₂O₂/O₃ process has a higher efficiency to degrade DNP, which shows that the COD removal efficiency can reach 39.7% (Fig. 2 a) only after 1 min treatment.

FIG. 3.

(a) UV spectra of DNP aqueous solution and (b) concentration variation of the generated NO₃⁻ and NO₂⁻ during the mFe⁰/H₂O₂/O₃ process under the optimal treatment conditions.

In addition, the generated NO₂⁻ and NO₃⁻ concentration variation during 15 min treatment by the mFe⁰/H₂O₂/O₃ process under the optimal treatment conditions was analyzed by IC. Figure 3b shows that NO₂⁻ concentration increased rapidly to the maximum (i.e., 37.4 mg/L) after 1 min treatment, and then, it gradually decreased to 5.6 mg/L after 15 min treatment. However, NO₃⁻ concentration gradually increased quickly to 216.4 mg/L after 15 min treatment. In addition, it can be calculated that the sum (i.e., 50.5 mg/L) of nitrate nitrogen (NO₃-N) and nitrite nitrogen (NO₂-N) in the effluent was lower than the theoretical nitrogen concentration (i.e., 76.5 mg/L) of 500 mg/L DNP aqueous solution. The results indicate that the main organic nitrogen of DNP was oxidized into NO₂⁻ and NO₃⁻, and the other organic nitrogen might be transferred into N₂, N₂O, or the smaller molecular organic nitrogen.

Possible reaction mechanism of mFe⁰/H₂O₂/O₃ process

SEM-EDS analysis of the mFe⁰ particles and generated flocculation

To investigate the possible reaction mechanism of mFe⁰/H₂O₂/O₃ process, the surface characteristics of fresh and reacted mFe⁰ particles and generated flocculation in mFe⁰/H₂O₂/O₃ process under optimal conditions were observed by SEM and EDS, respectively (Fig. 4). It could be observed from Fig. 4a and b that the morphology of fresh mFe⁰ particle was sponginess and smoothing, and only Fe was detected by EDS on the surface of fresh mFe⁰ particle. The results indicate that there was no iron oxide on the surface of fresh mFe⁰ particles. After the 15 min treatment by the mFe⁰/H₂O₂/O₃ process, it could be observed from Fig. 4c and d that two elements on the surface of mFe⁰ particles include Fe (77.39%, w/w) and O (22.61%, w/w), which suggests that some iron corrosion products deposited or adsorbed on the surface of mFe⁰ particles.

FIG. 4.

Scanning electron microscopy and energy-dispersive spectrometry spectra of (a, b) fresh and (c, d) reacted mFe⁰ particles and (e, f) generated flocculation in the mFe⁰/H₂O₂/O₃ process.

According to prior reports, the iron oxides could be composed of several materials (Fe₃O₄, Fe₂O₃, FeOOH, etc.), which could proceed the catalytic ozonation and Fenton-like reaction (Guo, 2010; Sui et al., 2010; Lv et al., 2012; Xu and Wang, 2012; Bing et al., 2015). In addition, the generated flocculation was also analyzed by SEM and EDS. From Fig. 4e and (f), it was observed that the major component elements of flocculation were O (40.3% w/w) and Fe (56.9% w/w), which suggests the generation of iron hydroxide.

XRD analysis

On the basis of SEM-EDS analysis, the fresh and reacted mFe⁰ particles and generated flocculation in the mFe⁰/H₂O₂/O₃ process were further analyzed by XRD. Figure 5 shows the XRD patterns of fresh and reacted mFe⁰ particles and generated flocculation. It can be observed from Fig. 5 that the reacted mFe⁰ particles after 15 min treatment show similar XRD patterns as that of fresh mFe⁰ particles. No visible crystalline peak was detected for iron oxide in all lines, which indicates the iron oxide on the surface of mFe⁰ is amorphous. In addition, the peak intensity was weak after the reaction of mFe⁰ particles, which may be due to iron oxide formed on its surface. Since Fe⁰ is a kind of moderate reduction metal, O₃/O₂ might react with Fe⁰ and resulted in the formation of iron oxide on the surface of Fe⁰ particles.

FIG. 5.

X-ray diffraction pattern of fresh and reacted mFe⁰ particles by the mFe⁰/H₂O₂/O₃ process.

XPS analysis

To further investigate the material composition on the surface of reacted mFe⁰ particles, the chemical state of Fe and O of mFe⁰ particles after the mFe⁰/H₂O₂/O₃ treatment was analyzed by XPS, and the results are shown in Fig. 6. In Fig. 6a, the survey scan of mFe⁰ particles reveals the presence of Fe 2p, O 1s, and C 1s. C1s peak (284.6 eV) was used to calibrate the acquired spectra. All XPS core level spectra were fitted using the Shirley background.

FIG. 6.

X-ray photoelectron spectroscopy spectra of reacted mFe⁰ particles by the mFe⁰/H₂O₂/O₃ process: (a) full-scale spectra, (b) Fe 2p spectra, and (c) O 1s spectra.

Figure 6b shows the XPS survey scan of Fe 2p in reacted mFe⁰ particles. The binding energies at 710.0–720.0 eV and 720.0–735.0 eV regions can be attributed to Fe 2p3/2 and Fe 2p1/2, respectively. The high-intensity binding peaks at 710.8 and 724.2 eV correspond to Fe³⁺ state of iron oxides (Zhang et al., 2012; Ji et al., 2017). Thus, the observed Fe 2p XPS spectra proved the existence of Fe₂O₃ generated on the surface of reacted mFe⁰ particles. In addition, the high-resolution scan of the O 1s core level is depicted in Fig. 6c. The spectrum shows the presence of two peaks at 529.7 and 531.4 eV. According to the literatures, the peak at 529.7–530.1 eV corresponds to O²⁻, and the peak at 531.1–531.7 eV can be attributed to surface-adsorbed oxygen species (i.e., O₂²⁻, O⁻, OH⁻) (Luo et al., 2008; Guo et al., 2010).

Mechanism

According to above analysis data and the present literatures, possible reaction mechanisms and synergistic effects of the mFe⁰/H₂O₂/O₃ process were illustrated in the following sections (Fig. 7).

FIG. 7.

Proposed reaction mechanism of mFe⁰/H₂O₂/O₃ process.

Reaction mechanisms between Fe⁰ and O₃

Since the standard oxidation reduction potential of O₃ can reach up to 2.07 mv, Fe⁰ might be directly oxidized by O₃ and generated Fe²⁺/Fe³⁺ [Eqs. (1) and (2)] (Xiong et al., 2016a). In addition, in acidic environment, H⁺ ions also would accelerate remarkably the iron corrosion rate and generated a plenty of Fe²⁺/Fe³⁺. Moreover, the DNP in aqueous solution was easy to be oxidized by single O₃ process (Fig. 2a) and transformed to small-molecule organic acids, which would also accelerate the iron corrosion rate. The generation of Fe²⁺/Fe³⁺ can effectively catalyze ozonation and generate a plenty of radicals to degrade pollutants [Eqs. (2 )–(6)].

Meanwhile, according to the analysis results of SEM-EDS, XRD, and XPS, it was proposed that iron oxides (i.e., Fe₂O₃) were generated on the surface of mFe⁰ particles. It is well known that Fe₂O₃ also could catalyze the decomposition of O₃ and generate the radicals to degrade pollutants. For example, Trapido et al. (2005) reported that Fe₂O₃ can effectively catalyze ozonation m-DNB at pH 3.0. Beltran et al. (2005) also found that the ozonation of oxalic acid in aqueous solution using Fe₂O₃/Al₂O₃ as catalyst can significantly improve the degradation efficiency of oxalic acid. Therefore, a typical mutual promotion did exist between O₃ and Fe⁰.

Reaction mechanisms between Fe⁰ and H₂O₂

The generated corrosion products of Fe⁰ (e.g., Fe²⁺ and Fe₂O₃) can take part in the Fenton or Fenton-like reaction to generate a plenty of HO^•, enhancing the treatment efficiency of mFe⁰/H₂O₂/O₃ process (Guo et al., 2010; Bokare and Choi, 2014). In addition, the H₂O₂ could also be generated by the reaction between O₂ and Fe⁰ in the acidic environment, and then, it reacts with Fe²⁺ and generates HO^•, which can nonselectively oxidize or mineralize the organic pollutants [Eqs. (7)–(8)] (Wang et al., 2010b). Moreover, a part of nascent Fe³⁺ can react with Fe⁰ to generate Fe²⁺, which take part in the Fenton reaction [Eq. (10)], thus further improve the treatment efficiency of mFe⁰/H₂O₂/O₃ process.

Reaction mechanisms between O₃ and H₂O₂

The reaction of O₃ and H₂O₂ can accelerate the decomposition of O₃ and generate radicals to degrade pollutants. The reaction equations are listed in Equations (11 )–(16) (Safarzadeh-Amiri, 2001; Medellin-Castillo et al., 2013). \documentclass{aastex}\usepackage{amsbsy}\usepackage{amsfonts}\usepackage{amssymb}\usepackage{bm}\usepackage{mathrsfs}\usepackage{pifont}\usepackage{stmaryrd}\usepackage{textcomp}\usepackage{portland, xspace}\usepackage{amsmath, amsxtra}\usepackage{upgreek}\pagestyle{empty}\DeclareMathSizes{10}{9}{7}{6}\begin{document} \begin{align*} {{ \rm{H}}_{ \rm{2}}}{{ \rm{O}}_{ \rm{2}}} \to { \rm{HO}}_{ \rm{2}}^{ \rm{ - }} + { \rm{ H}}_{}^{ \rm{ + }} \tag{11} \end{align*} \end{document} \documentclass{aastex}\usepackage{amsbsy}\usepackage{amsfonts}\usepackage{amssymb}\usepackage{bm}\usepackage{mathrsfs}\usepackage{pifont}\usepackage{stmaryrd}\usepackage{textcomp}\usepackage{portland, xspace}\usepackage{amsmath, amsxtra}\usepackage{upgreek}\pagestyle{empty}\DeclareMathSizes{10}{9}{7}{6}\begin{document} \begin{align*} {{ \rm{O}}_{ \rm{3}}}{ \rm{ + HO}}_{ \rm{2}}^{ \rm{ - }} \to { \rm{HO}}_{ \rm{2}}^ \bullet + { \rm{O}}{_{ \rm{3}}^ {\bullet { \rm{ - }}}} \tag{12} \end{align*} \end{document} \documentclass{aastex}\usepackage{amsbsy}\usepackage{amsfonts}\usepackage{amssymb}\usepackage{bm}\usepackage{mathrsfs}\usepackage{pifont}\usepackage{stmaryrd}\usepackage{textcomp}\usepackage{portland, xspace}\usepackage{amsmath, amsxtra}\usepackage{upgreek}\pagestyle{empty}\DeclareMathSizes{10}{9}{7}{6}\begin{document} \begin{align*} { \rm{HO}}_{ \rm{2}}^ \bullet \leftrightarrow { \rm{H}}_{}^{ \rm{ + }} + { \rm{O}}{_{ \rm{2}}^ {\bullet-} }\tag{13} \end{align*} \end{document} \documentclass{aastex}\usepackage{amsbsy}\usepackage{amsfonts}\usepackage{amssymb}\usepackage{bm}\usepackage{mathrsfs}\usepackage{pifont}\usepackage{stmaryrd}\usepackage{textcomp}\usepackage{portland, xspace}\usepackage{amsmath, amsxtra}\usepackage{upgreek}\pagestyle{empty}\DeclareMathSizes{10}{9}{7}{6}\begin{document} \begin{align*} { \rm{O}}{_{ \rm{2}}^ {\bullet { \rm{ - }}}}{ \rm{ + }}{{ \rm{O}}_{ \rm{3}}} \to { \rm{O}}_{ \rm{2}}^{}{ \rm{ + O}}{_{ \rm{3}}^ {\bullet { \rm{ - }}}} \tag{14} \end{align*} \end{document} \documentclass{aastex}\usepackage{amsbsy}\usepackage{amsfonts}\usepackage{amssymb}\usepackage{bm}\usepackage{mathrsfs}\usepackage{pifont}\usepackage{stmaryrd}\usepackage{textcomp}\usepackage{portland, xspace}\usepackage{amsmath, amsxtra}\usepackage{upgreek}\pagestyle{empty}\DeclareMathSizes{10}{9}{7}{6}\begin{document} \begin{align*} { \rm{O}}{_{ \rm{3}}^ {\bullet { \rm{ - }}}}{ \rm{ + }}{{ \rm{H}}^{ \rm{ + }}} \to { \rm{HO}}_{ \rm{3}}^ \bullet \tag{15} \end{align*} \end{document} \documentclass{aastex}\usepackage{amsbsy}\usepackage{amsfonts}\usepackage{amssymb}\usepackage{bm}\usepackage{mathrsfs}\usepackage{pifont}\usepackage{stmaryrd}\usepackage{textcomp}\usepackage{portland, xspace}\usepackage{amsmath, amsxtra}\usepackage{upgreek}\pagestyle{empty}\DeclareMathSizes{10}{9}{7}{6}\begin{document} \begin{align*} { \rm{HO}}_3^ \bullet \to { \rm{O}}_{ \rm{2}}^{}{ \rm{ + H}}{{ \rm{O}}^ \bullet } \tag{16} \end{align*} \end{document}

Reusability of mFe⁰ in the reaction system

To evaluate the reuse potential of mFe⁰ in the mFe⁰/H₂O₂/O₃ process, the reusability of mFe⁰ was investigated by conducting five successive experiments under the same optimal conditions. At the end of each cycle, the mFe⁰ powders were recovered through magnetic separation and drying, and then, they were reused with compensation of the mass loss for the next cycle. As shown in Fig. 8, the COD removal efficiencies were all above 78% in each cycle, and no obvious decline was observed after five cycles, which manifest the approving recyclability of mFe⁰.

FIG. 8.

COD removal for five successive mFe⁰/H₂O₂/O₃ process.

Conclusion

In this study, the mFe⁰/H₂O₂/O₃ process was developed to degrade DNP in aqueous solution. First, the optimal operation parameters (i.e., initial pH of 2.5, mFe⁰ dosage of 3.0 g/L, O₃ flow rate of 0.3 L/min, and H₂O₂ dosage of 25 mmol/L) of the mFe⁰/H₂O₂/O₃ process were observed by semibatch experiments. In this condition, COD removal efficiency reached 80.2% and was much higher than those of the mFe⁰/O₃ process (54.4%), mFe⁰/H₂O₂/O₂ process (52.6%), mFe⁰/H₂O₂ process (44.6%), O₃/H₂O₂ process (36.8%), sole O₃ process (23.3%), sole mFe⁰ process (9.0%), and sole H₂O₂ process (0%). The results indicate that high treatment efficiency of the mFe⁰/H₂O₂/O₃ process was mainly attributed to the synergetic effects between mFe⁰, H₂O₂, and O₃. Meanwhile, the B/C ratio was elevated from 0.25 to 0.61 after 15 min treatment by the mFe⁰/H₂O₂/O₃ process, which indicated the significant improvement of biodegradability. According to the literatures and analysis data of SEM-EDS, XRD, and XPS, finally, the possible reaction mechanisms of the mFe⁰/H₂O₂/O₃ process was proposed, which consist of four parts: (i) catalytic ozonation by Fe⁰ and its corrosion products, (ii) Fenton and Fenton-like reaction to degrade pollutants, (iii) direct oxidation by sole O₃ process, and (iv) the generated radicals by O₃/H₂O₂ process to degrade pollutants. Therefore, the mFe⁰/H₂O₂/O₃ process can be considered as an effective, robust, and feasible treatment method to degrade DNP in aqueous solution.

Footnotes

Author Disclosure Statement

No competing financial interests exist.

References

Bautitz

I.R.

, Velosa

A.C.

, and Nogueira

R.F.

(2012). Zero valent iron mediated degradation of the pharmaceutical diazepam. Chemosphere, 88, 688.

Beltrán

F.J.

, Rivas

F.J.

, and Montero-de-Espinosa

(2005). Iron type catalysts for the ozonation of oxalic acid in water. Water Res. 39, 3553.

Bing

, Hu

, Nie

, Yang

, and Qu

(2015). Mechanism of catalytic ozonation in Fe₂O₃/Al₂O₃@SBA-15 aqueous suspension for destruction of ibuprofen. Environ. Sci. Technol., 49, 1690.

Bokare

A.D.

, and Choi

(2014). Review of iron-free Fenton-like systems for activating H₂O₂ in advanced oxidation processes. J. Hazard. Mater., 275, 121.

Cai

, Liu

, Wang

, Zhang

, Zeng

, Yuan

, Ma

, Dong

, Liu

, and Luo

(2017). Silver phosphate-based Z-Scheme photocatalytic system with superior sunlight photocatalytic activities and anti-photocorrosion performance. Appl. Catal. B Environ., 208, 1.

Cao

X.Y.

, and Shiraishi

(2010). A mechanism of photocatalytic and adsorptive treatment of 2,4-dinitrophenol on a porous thin film of TiO₂ covering granular activated carbon particles. Chem. Eng. J., 160, 651.

Chen

, Dai

, Shi

, Li

, Wang

, and Yang

(2016). Sonocatalytic degradation of Rhodamine B catalyzed by β-Bi₂O₃ particles under ultrasonic irradiation. Ultrason. Sonochem., 29, 172.

Chen

, Liu

, Xia

, and Wang

(2017). Popcorn balls-like ZnFe 2 O 4 -ZrO 2 microsphere for photocatalytic degradation of 2,4-dinitrophenol. Appl. Surf. Sci., 407, 470.

Fukuchi

, Nishimoto

, Fukushima

, and Zhu

(2014). Effects of reducing agents on the degradation of 2,4,6-tribromophenol in a heterogeneous Fenton-like system with an iron-loaded natural zeolite. Appl. Catal. B Environ., 147, 411.

10.

Gong

, Lee

C.-S.

, Chang

Y.-Y.

, and Chang

Y.-S.

(2015). Novel self-assembled bimetallic structure of Bi/Fe0: The oxidative and reductive degradation of hexahydro-1,3,5-trinitro-1,3,5-triazine (RDX). J. Hazard. Mater., 286, 107.

11.

Guo

(2010). S-doped alpha-Fe2O3 as a highly active heterogeneous Fenton-like catalyst towards the degradation of acid orange 7 and phenol. Appl Catalysis B Environ. 96, 162.

12.

Guo

, Chen

, Fan

, Cai

, and Zhang

(2010). S-doped α-Fe2O3 as a highly active heterogeneous Fenton-like catalyst towards the degradation of acid orange 7 and phenol. Appl. Catal. B Environ., 96, 162.

13.

Hussain

, Zhang

, Huang

, and Du

(2012). Degradation of p-chloroaniline by persulfate activated with zero-valent iron. Chem. Eng. J., 203, 269.

14.

, Yin

, Zhang

, and Lai

(2018). Treatment of military primary explosives wastewater containing lead styphnate (LS) and lead azide (LA) by mFe 0 -PS-O 3 process. J. Cleaner Prod., 188, 860.

15.

, Zhang

, Li

, and Lai

(2017). Treatment of reverse osmosis (RO) concentrate from an old landfill site by Fe0/PS/O3 process. J. Chem. Technol. Biotechnol., 92, 2616.

16.

Kasprzyk-Hordern

, Ziółek

, and Nawrocki

(2003). Catalytic ozonation and methods of enhancing molecular ozone reactions in water treatment. Appl. Catal. B Environ., 46, 639.

17.

Lee

, Park

, and Harvey

OR.

(2013). Reduction of Chromium(VI) mediated by zero-valent magnesium under neutral pH conditions. Water Res. 47, 1136.

18.

Liu

, Wang

, and Li

X.-Z.

(2013). Electron efficiency of zero-valent iron for groundwater remediation and wastewater treatment. Chem. Eng. J. 215–216, 90.

19.

Luo

, Ai

, and Zhang

(2008). Fe@Fe₂O₃ Core−Shell Nanowires as Iron Reagent. 4. Sono−Fenton degradation of pentachlorophenol and the mechanism analysis. J. Phys. Chem. C, 112, 8675.

20.

, Hu

, Nie

, and Qu

(2012). Catalytic ozonation of toxic pollutants over magnetic cobalt-doped Fe₃O₄ suspensions. Appl. Catal. B Environ. 117–118, 246.

21.

, Zhang

L.-Z.

, Wang

Y.-H.

, Lei

S.-L.

, Luo

X.-B.

, Chen

S.-H.

, Zeng

G.-S.

, Zou

J.-P.

, Luo

S.-L.

, and Au

C.-T.

(2014). Mechanism of 2,4-dinitrophenol photocatalytic degradation by ζ-Bi₂O₃/Bi₂MoO₆ composites under solar and visible light irradiation. Chem. Eng. J., 251, 371.

22.

Martins

R.C.

, Lopes

D.V.

, Quina

M.J.

, and Quinta-Ferreira

R.M.

(2012). Treatment improvement of urban landfill leachates by Fenton-like process using ZVI. Chem Eng. J., 192, 219.

23.

Medellin-Castillo

N.A.

, Ocampo-Pérez

, Leyva-Ramos

, Sanchez-Polo

, Rivera-Utrilla

, and Méndez-Díaz

J.D.

(2013). Removal of diethyl phthalate from water solution by adsorption, photo-oxidation, ozonation and advanced oxidation process (UV/H₂O₂, O₃/H₂O₂ and O₃/activated carbon). Sci. Total Environ., 442, 26.

24.

Nawrocki

, and Kasprzyk-Hordern

(2010). The efficiency and mechanisms of catalytic ozonation. Appl. Catal. B Environ., 99, 27.

25.

W.-D.

, Dong

, and Lim

T.-T.

(2016). Generation of sulfate radical through heterogeneous catalysis for organic contaminants removal: Current development, challenges and prospects. Appl. Catalysis B. Environ., 194, 169.

26.

Pan

, Luo

, Fan

J.-J.

, Liu

D.-C.

, and Fu

(2012). Degradation of disperse Blue E-4R in aqueous solution by zero-valent iron/ozone. Clean Soil Air Water, 40, 422.

27.

Pillai

I.M.S.

, and Gupta

A.K.

(2015). Batch and continuous flow anodic oxidation of 2,4-dinitrophenol: Modeling, degradation pathway and toxicity. J Electroanal. Chem., 756, 108.

28.

, Mao

, Lv

, Sun

, Wang

, Yang

, and Xie

Y.F.

(2016). Pathway fraction of bromate formation during O3 and O3/H2O2 processes in drinking water treatment. Chemosphere, 144, 2436.

29.

Remya

, and Lin

J.-G.

(2011). Microwave-assisted carbofuran degradation in the presence of GAC, ZVI and H₂O₂: Influence of reaction temperature and pH. Sep. Purif. Technol., 76, 244.

30.

Roth

J.A.

, and Sullivan

D.E.

(2002). Solubility of ozone in water. Ind. Eng. Chem. Fundam., 20, 137.

31.

Safarzadeh-Amiri

(2001). O3/H2O2 treatment of methyl-tert-butyl ether (MTBE) in contaminated waters. Water Res. 35, 3706.

32.

Segura

, Martínez

, Melero

J.A.

, Molina

, Chand

, and Bremner

D.H.

(2012). Enhancement of the advanced Fenton process (Fe0/H2O2) by ultrasound for the mineralization of phenol. Appl. Catal. B Environ., 113, 100.

33.

Sellers

R.M.

(1980). Spectrophotometric determination of hydrogen peroxide using potassium titanium(IV) oxalate. Analyst, 105, 950.

34.

Shiraishi

, Miyawaki

, and Chand

(2015). A mechanism of the photocatalytic decomposition of 2,4-dinitrophenol on TiO2 immobilized on a glass surface. Chem. Eng. J., 262, 831.

35.

Son

H.S.

, Im

J.K.

, and Zoh

K.D.

(2009). A Fenton-like degradation mechanism for 1,4-dioxane using zero-valent iron (Fe0) and UV light. Water Res. 43, 1457.

36.

Song

, Zhou

, Liu

, Xie

G.-J.

, Wang

, Zhang

, Liu

, Zhou

, and Wang

(2016). Improving dewaterability of anaerobically digested sludge by combination of persulfate and zero valent iron. Chem. Eng. J., 295, 436.

37.

Srithep

, and Phattarapattamawong

(2017). Kinetic removal of haloacetonitrile precursors by photo-based advanced oxidation processes (UV/H₂O₂, UV/O₃, and UV/H₂O₂/O₃). Chemosphere, 176, 25.

38.

Sui

, Sheng

, Lu

, and Tian

(2010). FeOOH catalytic ozonation of oxalic acid and the effect of phosphate binding on its catalytic activity. Appl. Catal. B Environ., 96, 94.

39.

Trapido

, Veressinina

, Munter

, and Kallas

(2005). Catalytic ozonation of m-dinitrobenzene. Ozone Sci. Eng., 27, 359.

40.

Wang

, Wang

H.-L.

, Jiang

W.-F.

, and Li

Z.-Q.

(2009). Photocatalytic degradation of 2,4-dinitrophenol (DNP) by multi-walled carbon nanotubes (MWCNTs)/TiO2 composite in aqueous solution under solar irradiation. Water Res. 43, 204.

41.

Wang

H.-L.

, Liang

W.-Z.

, Zhang

, and Jiang

W.-F.

(2010a). Solar-light-assisted Fenton oxidation of 2,4-dinitrophenol (DNP) using Al₂O₃-supported Fe(III)-5-sulfosalicylic acid (ssal) complex as catalyst. Chem. Eng. J., 164, 115.

42.

Wang

, and Bai

(2017). Fe-based catalysts for heterogeneous catalytic ozonation of emerging contaminants in water and wastewater. Chem. Eng. J., 312, 79.

43.

Wang

K.S.

, Lin

C.L.

, Wei

M.C.

, Liang

H.H.

, Li

H.C.

, Chang

C.H.

, Fang

Y.T.

, and Chang

S.H.

(2010b). Effects of dissolved oxygen on dye removal by zero-valent iron. J. Hazard. Mater., 182, 886.

44.

Wei

, Song

, Tu

, Zhao

, and Zhi

(2013). Pretreatment of dry-spun acrylic fiber manufacturing wastewater by Fenton process: Optimization, kinetics and mechanisms. Chem. Eng. J., 218, 319.

45.

Xiong

, Lai

, Yang

, Zhou

, Wang

, and Fang

(2015). Comparative study on the reactivity of Fe/Cu bimetallic particles and zero valent iron (ZVI) under different conditions of N 2, air or without aeration. J. Hazard. Mater, 297, 261.

46.

Xiong

, Lai

, Yuan

, Cao

, Yang

, and Zhou

(2016a). Degradation of p-nitrophenol (PNP) in aqueous solution by a micro-size Fe0/O3 process (mFe0/O3): Optimization, kinetic, performance and mechanism. Chem. Eng. J., 302, 137.

47.

Xiong

, Yuan

, Lai

, Yang

, and Zhou

(2016b). Mineralization of ammunition wastewater by a micron-size Fe0/O₃ process (mFe0/O₃). RSC Adv. 6, 55726.

48.

, and Wang

(2012). Fenton-like degradation of 2,4-dichlorophenol using Fe₃O₄ magnetic nanoparticles. Appl. Catal. B Environ. 123–124, 117.

49.

Yoon

I.H.

, Kim

K.W.

, Bang

, and Kim

M.G.

(2011). Reduction and adsorption mechanisms of selenate by zero-valent iron and related iron corrosion. Appl. Catal. B Environ., 104, 185.

50.

Zeng

, Zou

, Li

, Sun

, Chen

, and Shao

(2012). Ozonation of phenol with O₃/Fe(II) in acidic environment in a Rotating Packed Bed. Ind. Eng. Chem. Res., 51, 10509.

51.

Zhang

, Zhang

, Wang

, Liu

, Zeng

, Xia

, Liu

, and Luo

(2017a). Facile synthesis of bird's nest-like TiO₂ microstructure with exposed (001) facets for photocatalytic degradation of methylene blue. Appl. Surf. Sci., 391, 228.

52.

Zhang

, Ji

, Zhang

, Pan

, and Lai

(2017b). Catalytic ozonation of N, N-dimethylacetamide(DMAC) in aqueous solution using nanoscaled magnetic CuFe₂O₄. Sep. Purif. Technol., 193, 368.

53.

Zhang

, Xiong

, Ji

, Lai

, and Yang

(2017c). Pretreatment of shale gas drilling flowback fluid (SGDF) by the microscale Fe0/Persulfate/O₃ process (mFe0/PS/O₃). Chemosphere. 176, 192.

54.

Zhang

, Wang

, Liu

, Ding

, Zhang

, Zeng

, Liu

, and Luo

(2016). A bamboo-inspired hierarchical nanoarchitecture of Ag/CuO/TiO(2) nanotube array for highly photocatalytic degradation of 2,4-dinitrophenol. J. Hazard. Mater., 313, 244.

55.

Zhang

, Li

, Sheng

, Zhang

, and Zheng

(2012). Enhanced Cr(VI) removal by using the mixture of pillared bentonite and zero-valent iron. Chem. Eng. J. 185–186, 243.

56.

Zhao

, Dong

, Wang

, Chen

, Wang

, Liu

, Gao

, and Zhou

(2017). Advanced oxidation removal of hypophosphite by O₃/H₂O₂ combined with sequential Fe(II) catalytic process. Chemosphere, 180, 48.

57.

Zhu

, Tian

, Liu

, and Chen

(2011). Optimization of Fenton and electro-Fenton oxidation of biologically treated coking wastewater using response surface methodology. Sep. Purif. Technol., 81, 444.

Degradation of 2,4-Dinitrophenol in Aqueous Solution by Microscale Fe 0 /H 2 O 2 /O 3 Process

Abstract

Abstract

Introduction

Materials and Methods

Chemicals

Experimental procedure

Analytical method

Results and Discussion

Optimization of operation parameters

Effect of initial solution pH

Effect of mFe0 dosage

Effect of O3 flow rate

Effect of H2O2 dosage

Control experiments

COD removal during the treatment process

Biodegradability

Total iron ion concentration

The utilization of H2O2

UV-Vis spectral analysis and transformation of nitrogen

Possible reaction mechanism of mFe0/H2O2/O3 process

SEM-EDS analysis of the mFe0 particles and generated flocculation

XRD analysis

XPS analysis

Mechanism

Reaction mechanisms between Fe0 and O3

Reaction mechanisms between Fe0 and H2O2

Reaction mechanisms between O3 and H2O2

Reusability of mFe0 in the reaction system

Conclusion

Footnotes

Author Disclosure Statement

References

Effect of mFe⁰ dosage

Effect of O₃ flow rate

Effect of H₂O₂ dosage

The utilization of H₂O₂

Possible reaction mechanism of mFe⁰/H₂O₂/O₃ process

SEM-EDS analysis of the mFe⁰ particles and generated flocculation

Reaction mechanisms between Fe⁰ and O₃

Reaction mechanisms between Fe⁰ and H₂O₂

Reaction mechanisms between O₃ and H₂O₂

Reusability of mFe⁰ in the reaction system