Introduction

Water pollution by phenolic pollutants is a serious environmental problem, and has generated widespread concern because of their low biodegradability, high biotoxicity, and strong cumulative effects (Zeng et al. 2013). Several advanced oxidation technologies have been applied to remove carcinogenic and harmful organic pollutants from water and wastewater, such as Fenton processes (Wang et al. 2018), electrochemical methods (Massa et al. 2018), photocatalytic technologies (Vaiano et al. 2018), and ozone oxidation (Zhao et al. 2013).

Ozone (O3) is an effective chemical oxidant (2.07 V) that can react with organic contaminants by directly oxidizing them with ozone or indirectly oxidizing them with hydroxyl radicals (·OH)(Wang and Bai 2017). Organic compounds containing unsaturated bonds or aromatic groups are particularly vulnerable to these types of oxidations (Zhu et al. 2017a). Ozonation alone is an effective oxidation process that has been used to oxidize contaminants in water and wastewater, decrease water color, control taste and odor, and kill microorganisms (Edwards-Brandt et al. 2007; Gümüş and Akbal 2017). However, O3 has a low solubility and stability in water and is a highly selective oxidant, making mineralization of organic compounds by ozonation alone is more difficult (Wang and Bai 2017). In view of this, catalytic ozonation has recently emerged as a popular technology to oxidize refractory organic pollutants because of its high efficiency, low cost, and simple application (Lv et al. 2010; Li et al. 2017).

Catalytic ozonation has been developed to enhance the production of ·OH (2.8 V) for the non-selective oxidation of organic pollutants (Ma et al. 2014), thus overcoming the disadvantages of ozonation alone. Both heterogeneous and homogeneous catalytic ozonation processes have been developed for water treatment (Wu et al. 2008; Wang and Bai 2017). However, compared with the possible secondary pollution caused by metallic ions in homogeneous catalytic ozonation processes, the heterogeneous catalytic ozonation is a more promising method (Nawrocki and Kasprzyk-Hordern 2010). The most widely used materials in heterogeneous catalytic ozonation are carbon-based materials (Rivera Utrilla et al. 2011), metal oxides (Cheng et al. 2017), metal oxides attached to support materials (Afzal et al. 2016), and natural minerals (Qi et al. 2012). Unfortunately, the synthesis of most catalysts is complex and expensive, which limits their industrial applications (Chen et al. 2014a). Heterogeneous catalysts based on cheap and stable materials have been proposed to overcome these shortcomings. One prominent example is granular activated carbon (GAC), which has been used in wastewater ozonation treatments due to its low cost, abundant reserves, ease of separation and recovery, and excellent mechanical properties (Álvarez et al. 2011; Ma et al. 2016). In recent decades, many studies have been published on GAC modification using supported transition metals and their oxides, such as Mn, Fe, Cu, Co, and Ce (Chen et al. 2014a; Huang et al. 2015; Li et al. 2017). The GAC-loaded metal oxides has been reported to possess more active sites and have higher catalytic activities (Chen et al. 2014b; Tang et al. 2016).

It has been previously reported that some non-noble metal oxides with variable oxidation states and fast electron transfer abilities, such as manganese and iron oxides, have been used as active materials because of their high catalytic activities and wide availability (Liu et al. 2016; Liang et al. 2016). Iron oxides have been used in a variety of applications, including the removal of refractory organic compounds such as 2, 4, 6-trichloroanisole (Qi et al. 2012), dibutyl phthalate (Huang et al. 2015), and atrazine (Zhu et al. 2017a). Iron oxides have been reported to form ·OH by redox recycling of Fe2+ and Fe3+ (Bai et al. 2016). In addition, manganese oxides have also been reported to be cheap and eco-friendly catalysts for the removal of organic compounds (Faria et al. 2009), and they exhibit significant catalytic performance in the degradation of p-chloronitrobenzene (Cheng et al. 2017). It is generally known that manganese has variable valences states from − 3 to + 7 because of its special electronic structure (Liu et al. 2016). Therefore, the manganese oxides can easily generate mobile electrons, which would increase the formation rate of ·OH by creating a mobile-electron environment during catalytic ozonation (Lv et al. 2010; Liu et al. 2016). Since the electron mobility of manganese oxides can be further promoted by incorporating other elements (Du et al. 2018), it is therefore hypothesized that an enhanced catalytic activity could be obtained by combining iron and manganese oxides. To the best of our knowledge, few studies have focused on heterogeneous catalytic ozonation performance using Fe- and/or Mn-loaded GACs as catalysts to remove phenol and actual secondary petrochemical effluent. Furthermore, the interactions between phenol and Fe- and/or Mn-loaded GACs in ozonation processes remain unknown.

Herein, a series of Fe and/or Mn/GAC catalysts were prepared, optimized and characterized. This paper is focused on the following three aspects: (i) evaluation of the catalytic degradation performance of phenol using three types of catalysts (Fe/GAC (0.10), Mn/GAC (0.10), and Fe-Mn/GAC (0.10:0.10)); (ii) elucidation of the catalytic ozonation mechanism and the phenol degradation pathway; (iii) verification of the application of Mn/GAC in the advanced treatment of secondary petrochemical effluent.

Experimental

Chemicals and samples

Coconut shell granular activated carbon was purchased from Sigma-Aldrich (Sigma-Aldrich Co. Ltd., USA). Phenol (> 99.0%) and standard samples of some intermediates were obtained from Sinopharm (Sinopharm Chemical Reagent Co., Ltd., China). Other reagents such as Mn (NO3)2, Fe (NO3)3·9H2O, Na2CO3, and tert-butanol were of analytical grade and used without further purification. The Vibrio fischeri used for the toxicity test was obtained from the Yangtze Delta Region Institute of Tsinghua University (Zhejiang, China). Secondary petrochemical effluent was collected from Tianshi (Tianshi Energy Co. Ltd., China), and the water quality parameters were as follows: 422.5 mg L−1 of COD, 100.0 mg L−1 of phenol (artificial preparation), 0.011 of BOD5/COD, 4.6 mg L−1 of NH4+-N, 148.9 mg L−1 of TOC, and solution pH of 8.3.

Fe and/or Mn/GAC catalysts were synthesized via impregnation-calcination using previously optimized conditions (Tang et al. 2016; Du et al. 2018). Initially, GAC (18–40 mesh) was washed with deionized water repeatedly and then dried at 105±1 °C for 24 h. Secondly, a fixed amount of GAC was mixed with 200 ml prepared metal salt solution (50 wt.% Mn (NO3)2, Fe (NO3)3·9H2O, and deionized water) in different molar concentrations at room temperature for 6 h. The mixture was then filtered, washed thoroughly with deionized water, and dried at 105±1 °C for 6 h. Finally, dried samples were calcined at 500 °C for 3 h at a heating rate of 10 °C/min, and they were then sieved for further experiments (18–40 mesh). The obtained catalysts were labeled with the molar concentration of Fe and Mn in the impregnated solution (e.g., when the molar concentration of Fe or Mn was 0.10 M, the catalysts were named Fe/GAC (0.10) and Mn/GAC (0.10)).

Characterization of samples

To study the basic characteristics of Fe and/or Mn/GAC catalysts, the proximate characterization analyses were performed by Chinese National Standards (GB/T 212-2008). The morphologies and sizes of the catalysts were examined using a scanning electron microscopy (SEM, SSX-550, Shimadzu, Japan) at 30 kV. The chemical composition was analyzed using an energy-dispersive spectrometer (EDS, SSX-550, Shimadzu, Japan). The EDS data was randomly obtained from three different catalyst regions. The crystalline structures of samples were measured by powder X-ray diffraction (XRD, D8 Focus, Bruker, Germany), and the 2θ data from 10 ̊ to 80 ̊ was collected at 10 ̊/min with a 0.07 ̊ step. Catalyst surface properties were measured by Fourier-transform infrared spectroscopy (FTIR, Vertex 70 V, Bruker, Germany) with a 4000–400 cm−1 spectral range. Chemical states of the constituent elements on the catalysts were evaluated by X-ray photoelectron spectroscopy analysis (XPS, ESCALAB250Xi, Thermo Fisher, USA). Electrochemical experiments were performed on a basic electrochemical system (CHI-660B, Chinehwa, China). All experiments were conducted using a three-electrode cell configuration with a catalyst-modified conductive glass (prepared by dip-coating and drying in air at 70±1 °C) as the working electrode, a platinum wire as the auxiliary electrode and a saturated calomel electrode as the reference electrode. In addition, the electrolyte was a 3 wt% NaCl solution.

Catalytic ozonation experiments

Catalytic ozonation of phenol and secondary petrochemical effluent by using Fe and/or Mn/GAC catalysts was carried out in a lab-scale reactor with an effective volume of 2.0 L (internal diameter = 8.0 cm and height = 50.0 cm). The reactor used bottom ventilation, which allowed the catalysts to mix more uniformly. A water sample (2.0 L; pH = 7.0) was added into the reactor, and then 2.0 g catalysts were introduced. Ozone generated from dried oxygen gas by a laboratory ozone generator (CF-G-3-10 g, QingDaoGuolin, China) was bubbled into the bottom of the reactor at a flow rate of 3.0 L/min by a titanium alloy porous diffuser. The ozone concentration was controlled by adjusting the electricity voltage, and the residual ozone in the off-gas was absorbed by a 20% KI solution. Ozonation experiments were terminated by bubbling a stream of N2 for 5 min. Used catalysts were separated and then dried at 105±1 °C for 24 h after being washed with deionized water to prepare them for further characterization and reuse.

Analytical methods

The concentrations of phenol and typical intermediate products were measured using high-performance liquid chromatography (HPLC, 1260 Series, Agilent, USA) with a VWD detector. The stationary phase was a mixture of acetonitrile and ultrapure water (7:3, by volume), and the injection volume was 20 μL. Intermediate products were analyzed using a gas chromatography coupled with mass spectrometry (GC-MS, 7860B/5977B Series, Agilent, USA). The injector temperature was 275 °C, and the temperature of the column oven was initially set at 70 °C for 5 min, at a ramp rate of 5 °C/min to 275 °C where it was held for 5 min. Helium was the carrier gas at a constant flow rate of 1 mL/min. Samples were extracted three times with hexane and then concentrated by purging with nitrogen gas. The concentrations of dissolved Fe and Mn in the treated wastewater during catalytic ozonation using Fe and/or Mn/GACs were analyzed by inductively coupled plasma optical emission spectrometry (ICP-OES, iCAP7000, Thermo, USA). The total organic carbon (TOC) was analyzed with a Multi TOC/TN Analyzer (2100, Analytik Jena, Germany). Ozone concentrations in the gaseous and aqueous phases were measured using the iodometric and indigo methods (Ozone Standards Committee Method), respectively. BOD5, COD, and NH4+-N were analyzed according to standard examination methods of water and wastewater (CSEPB 2002). The acute toxicity test was conducted using a microorganism fluorescence detector (AF-100, TOA-DKK, Japan) according to the standard method (UNE-EN-ISO 11348-32007) (Li et al. 2018).

Results and discussion

Characterization of catalysts

Catalysts were characterized using SEM, EDS, XRD, and FTIR (see Supplementary Material Test S1, Table S1, and Figs. S1-S2). As shown in Fig. S1, it is clear that Fe and Mn oxide nanoparticles were successfully and evenly coated on the surface of GACs. EDS analysis (Table S1) showed that the atomic percentage of Fe on the surfaces of Fe/GAC (0.10) and Fe-Mn/GAC (0.10:0.10) was ~ 0.90%, and that of Mn on the surfaces of Mn/GAC (0.10) and Fe-Mn/GAC (0.10:0.10) was ~ 0.30%. According to XRD analysis (Fig. S2), no diffraction Fe and Mn oxide diffraction peaks were observed on the surface of the catalysts, indicating the oxides were amorphous (Chen et al. 2017a; Li et al. 2017). FTIR spectroscopy (Fig. S3) showed a decrease in the O-H peak intensity after catalytic ozonation, demonstrating that the O-H might be partially consumed during the reaction (Bai et al. 2016). Moreover, the addition of Fe or Mn oxides did not increase the hydroxyl group density on the catalyst surfaces.

To better understand the roles of Fe and Mn oxides in a catalytic ozonation system, XPS spectra of Fe and Mn were recorded before and after catalytic ozonation of phenol. There were two main asymmetric peaks in the Fe2p spectra (Fig. 1a) at 711.0 ± 0.2 and 725.0 ± 0.2 eV which corresponded to Fe2p3/2 and Fe2p1/2, respectively. These positions are consistent with the iron assignments of magnetite (Tan et al. 2014; Zhu et al. 2017a), showing that Fe3+ from magnetite was present on the catalyst surfaces. Based on the Fe2p data, Fe2+ and Fe3+ accounted for 62.78% and 37.22% (Fex+/(Fe2+ + Fe3+)) in the fresh Fe-Mn/GAC catalyst, while 70.90% and 29.10% were in the used catalyst, respectively (Table S2). This verifies that Fe2+ and Fe3+ participated in the redox reaction during the catalytic ozonation of phenol. The Mn2p regions of the XPS spectra (Fig. 1b) have Mn2p3/2 binding energy peaks located at 641.35 eV, 642.80 eV, and 645.90 eV, which correspond to Mn2+ (MnO), Mn3+ (Mn2O3 or Mn3O4), and Mn4+ (MnO2), respectively (Umezawa and Reilley 1978; Strohmeier and Hercules 1984; Zhang et al. 2015; Liu et al. 2016; Du et al. 2018). Interestingly, as shown in Table S2, the percentages of different manganese ions (Mnx+/(Mn2+ + Mn3+ + Mn4+)) showed slight changes after the catalytic ozonation of phenol (Table S2), which indicated that the transitions occurred among Mn2+, Mn3+, and Mn4+ during the redox process. For example, the relative amount of Mn4+/(Mn2+ + Mn3+ + Mn4+) decreased from 23.87 to 19.85% due to the reaction between the strongly oxidizing Mn4+ and organic compounds (Li et al. 2015). As previously reported (Lv et al. 2010; Bai et al. 2016), Fe2+-Fe3+ and Mn2+-Mn3+-Mn4+ redox cycles play key roles in promoting the generation of ·OH. Therefore, Mn and Fe oxides on the catalyst surfaces have strong catalytic ozonation abilities.

Fig. 1
figure 1

Fe2p (a) and Mn 2p (b) XPS spectra of the fresh and used Fe or Mn/GAC (0.10) catalysts

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Catalytic ozonation performance of catalysts

Figure 2 shows the degradation and mineralization of phenol in adsorption, ozonation, and catalytic ozonation systems, as well as the changes in O3 concentration in solution. The removal performances of phenol and TOC with respect to time under different experimental conditions were evaluated as shown in Fig. 2a and b, respectively. In Fig. 2a, ozonation alone removed 88.8% of phenol within 25 min, while phenol was completely eliminated in the first 25 min using GAC/O3, Fe-Mn/GAC (0.10:0.10)/O3, Fe/GAC (0.10)/O3, and Mn/GAC (0.10)/O3. O3 was a powerful oxidant that reacted with organic compounds containing aromatic moieties and unsaturated bonds, which led to the rapid and complete removal of phenol (Zhu et al. 2017a). However, the catalytic ozonation systems using Fe and/or Mn/GACs had better phenol degradation performances than the GAC/O3 system. This indicated that adding Fe and Mn improved the system’s catalytic phenol ozonation ability.

Fig. 2
figure 2

Removal of phenol (a) and TOC (b) in aqueous dispersions of various catalysts with O3 (Cphenol = 200.0 mg L−1; catalysts dosage = 1.0 g L−1; O3 concentration = 22.0 mg L−1). Removal of TOC for adsorption of various catalysts (c) (Cphenol = 200.0 mg L−1; catalysts dosage = 1.0 g L−1). Residue ozone concentration in different systems (d) (Cphenol = 200.0 mg L−1; catalysts dosage = 1.0 g L−1; O3 concentration = 22.0 mg L−1)

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As shown in Fig. 2b, only 31.8% of TOC was removed at a reaction time of 90 min when using O3 alone. The degradation of phenol formed several intermediate products which were further converted to saturated compounds such as carboxylic acids, which could not be directly mineralized by O3 (Ramseier and Gunten 2009). Consequently, the degree of mineralization of phenol using O3 alone was not always satisfactory. For the GAC/O3 system, the TOC removal efficiency reached 47.0% due to the adsorption and slight catalytic ability of GAC (Ma et al. 2004). The Fe/GAC (0.10)/O3 and Fe-Mn/GAC (0.10:0.10)/O3 showed higher activities than GAC/O3, with TOC removal efficiencies of 56.4% and 60.1%, respectively. The Mn/GAC (0.10)/O3 system had the highest degradation activity, and the TOC concentration in solution decreased by 70.7%. Obviously, adding Fe and/or Mn on the surface of the GAC significantly enhanced the TOC removal activity of the catalysts. It is generally believed that metal-supported catalysts can accelerate conversion of O3 to ·OH free radicals and increase the O3 utilization efficiency (Li et al. 2017). It is this behavior which gave Fe/GAC (0.10), Mn/GAC (0.10), and Fe-Mn/GAC (0.10:0.10) their excellent catalytic performances.

Since adsorption processes are very important for the mineralization of organic compounds during heterogenous catalytic ozonation, the TOC adsorption of different catalysts was investigated. As shown in Fig. 2c, 16.7%, 18.8%, 20.4%, and 21.0% of TOC were removed by the GAC, Fe/GAC (0.10), Fe-Mn/GAC (0.10:0.10), and Mn/GAC (0.10) systems, respectively, over a 90-min reaction. This demonstrates that the catalyst adsorption played an important role in the TOC removal process. Compared with the GAC adsorption process, the Fe and/or Mn/GACs showed no obvious increases in the TOC removal during adsorption due to low metal loadings. This suggests that the TOC removal in solution was mainly accomplished by oxidation.

The adsorption and ozonation experiments showed that the TOC removal efficiency of Fe and/or Mn/GAC/O3 systems was higher compared with using adsorption or ozonation alone. This suggests that significant interactions occur between the catalysts and O3. To verify this hypothesis, different ozone concentrations were tested in different reaction systems, and the results are shown in Fig. 2d. The ozone concentration increased rapidly and then reached a dynamic equilibrium in all ozonation systems within the first 20 min. Moreover, the order of ozone equilibrium concentration was non-catalyst > GAC > Fe/GAC (0.10) > Fe-Mn/GAC (0.10:0.10) > Mn/GAC (0.10), which shows that adding catalysts significantly enhanced ozone decomposition. The order of ozone utilization efficiency was the same as the phenol mineralization efficiency order of the above five ozonation systems. Therefore, the catalytic performances of Fe and/or Mn/GACs were responsible for O3 decomposition during ozonation. These results are similar to previous studies (Zhu et al. 2017a; Chen et al. 2018) which indicated that catalysts play important roles in mineralizing phenol.

Effects of operating factors

The metal oxide loading affects not only the aggregation and dispersion of the catalyst, but also its crystallization and redox properties (Du et al. 2018). Therefore, the effect of Fe and/or Mn ion impregnation on phenol mineralization was evaluated, and the results are shown in Fig. 3. In a single Fe impregnation experiment, the removal efficiency of TOC showed a remarkable increase as the Fe concentration increased, reaching a maximum (56.4%) at an Fe impregnation concentration of 0.10 mol L−1, and then began to decline. Interestingly, the single Mn impregnation experiment showed a similar trend, and a maximum TOC removal efficiency of 70.7% was obtained at a Mn impregnation concentration of 0.10 mol L−1. Virgin GAC cannot provide enough active sites to catalyze the ozonation of phenol. However, the active sites of Fe or Mn/GACs increased as the Fe and Mn ion impregnation concentration increased, which enhanced the phenol mineralization capacities (He et al. 2010; Chen et al. 2018). On the other hand, overloading the GAC surface with Fe and Mn may block pores and decrease the total pore volume, specific surface area, and mean pore diameter of the catalysts, which would decrease the catalyst active sites (Li et al. 2009). Consistent with the above results, Dai et al. (2014) found that excess Ce loading decreased the catalytic ozonation performance of p-TSA by Ce/AC due to a decrease in the pore volume, special surface area, and pore diameter of the catalyst.

Fig. 3
figure 3

Effect of Fe or/and Mn impregnation concentrations on mineralization of phenol (Cphenol = 200.0 mg L−1; catalysts dosage = 1.0 g L−1; ozone concentration = 22.0 mg L−1)

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To investigate the catalytic properties of catalysts with Fe-Mn composite metal oxides, the efficiency of phenol mineralization was assessed using catalysts with different Fe-Mn impregnation ratios. As shown in Fig. 3, TOC removal by Fe-Mn/GAC (0.50:0.10) and Fe-Mn/GAC (0.10:0.10) was higher than that of other ratios, but far lower than that of Mn/GAC (0.10). These results show that 0.10 mol L−1 was the threshold for inhibiting the catalytic ozonation performance of Fe or Mn/GAC. Therefore, there was a decrease in the number of active sites in Fe-Mn/GAC (0.10:0.10) due to pore blockage, which reduced the catalytic performance. In addition, it was found that the mineralization of phenol followed a first-order kinetic model in the Fe and/or Mn/GAC system, as shown in Fig. 3. The rate constant order of these ozonation systems was identical to the order of phenol mineralization efficiency. Moreover, the rate constant reached a maximum (2.0 L mg−1 min−1) at a Mn impregnation concentration of 0.10 mol L−1, indicating that it had superior catalytic performance.

To investigate the roles of ·OH in ozonation alone and catalytic ozonation systems, respectively, TBA was selected as a scavenger, because of its high reaction rate constant of 6 × 108 M−1 s−1 with ·OH and only 3 × 10−3 M−1 s−1 with ozone (Buxton et al. 1988). Figure 4a shows that the phenol removal efficiency reaches 100% in the catalytic ozonation system and 85.7% in ozonation alone system within the first 40 min. However, when TBA was added, the phenol removal efficiency decreased by 20.0% and 10.7% in the catalytic ozonation system with Mn/GAC (0.10) and ozonation alone at 40 min, respectively. Compared to the ozonation alone, a 9.3% decrease was obtained in the catalytic ozonation system in the presence of TBA, suggesting that catalytic ozonation by Mn/GAC (0.10) generated more ·OH.

Fig. 4
figure 4

Effect of TBA in phenol catalytic ozonation process (a) (Cphenol = 200.0 mg L−1; Mn/GAC (0.10) dosage = 1.0 g L−1; ozone concentration = 11.18 mg L−1; TBA = 1.0 g L−1). TOC removal in the presence of carbonate ions (b) (Cphenol = 200.0 mg L−1; Mn/GAC (0.10) dosage = 1.0 g L−1; ozone concentration = 22.0 mg L−1)

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It is well known that carbonate and bicarbonate are ozone decomposition inhibitors by scavenging ·OH without generating \( {\mathrm{O}}_2^{\cdotp -} \) or other free radicals that can accelerate ozone decomposition (Staehelin and Hoigné 1985). Thus, the effect of carbonate on phenol mineralization was investigated by adding 250 mg L−1 and 500 mg L−1 Na2CO3 to the substrate solution. As depicted in Fig. 4b, the phenol mineralization was significantly affected after 250 mg L−1 Na2CO3 was added, and the removal efficiency of TOC decreased by 18.1%. Further increasing the concentration of Na2CO3 to 500 mg L−1 did not significantly inhibit phenol mineralization. The reactions between ·OH and CO32−/HCO3− can be described using Eqs. (1) and (2) (Buxton et al. 1988):

$$ {\mathrm{HCO}}_3^{-}+\cdotp \mathrm{OH}\to \kern0.5em {\mathrm{HCO}}_3^{\cdotp -}+{\mathrm{OH}}^{-}\kern0.5em k=8.5\times {10}^6{\mathrm{M}}^{-1}{\mathrm{s}}^{-1} $$
(1)
$$ {\mathrm{CO}}_3^{2-}+\cdotp \mathrm{OH}\to {\mathrm{CO}}_3^{\cdotp -}+{\mathrm{OH}}^{-}\kern0.5em k=3.9\times {10}^8{\mathrm{M}}^{-1}{\mathrm{s}}^{-1} $$
(2)

The result obtained from the above experiment indicated that carbonate significantly decreased the TOC removal efficiency, which further verified that ·OH played an important role in catalytic ozonation using Mn/GAC (0.10).

Catalytical ozonation mechanism

Identification of intermediate products and their degradation pathways

GC-MS and HPLC showed that many intermediate products formed during the catalytic ozonation process, as shown in Table S2. According to the formation of intermediate products during the phenol degradation, possible pathways were tentatively proposed and are depicted in Fig. 5. It can be seen that the phenol oxidation during catalytic ozonation proceeded initially through the hydroxylation of the aromatic ring to form catechol, hydroquinone, and resorcinol. The dihydroxybenzenes, such as catechol and hydroquinone, were in redox equilibrium with benzoquinones (Zazo et al. 2005). Dihydroxybenzenes were further oxidized via hydroxylation to form trihydroxybenzenes (benzenetriol) (Kermani et al. 2018). These were decomposed into muconic, succinic, maleic, and fumaric acids via benzene ring cleavage and were further decomposed into oxalic and acetic acids (Zazo et al. 2005; Ramseier and Gunten 2009; Dai et al. 2014). Finally, aliphatic products containing carbonyl and carboxyl groups were oxidized to CO2 and H2O. The above phenol degradation pathway was mainly ascribed to two types of oxidation in the Fe and/or Mn/GAC/O3 systems. In one type, direct oxidation of ozone was extremely selective for unsaturated aromatic and aliphatic compounds as well as specific functional groups via Criegee mechanism (Criegee 1953; Turhan and Uzman 2008). In the other type, the indirect oxidation of ·OH played an important role in the cleavage of benzene rings and unsaturated bonds due to its non-selectivity and high oxidation ability.

Fig. 5
figure 5

Proposed pathway of phenol degradation by catalytic ozonation (Cphenol = 200.0 mg L−1; catalysts dosage = 1.0 g L−1; ozone concentration = 22.0 mg L−1)

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Exploration of the catalytic mechanism

During heterogeneous catalytic ozonation, ozone can be efficiently converted to free radicals through its interactions with a catalyst (Zhu et al. 2017b; Centurião et al. 2019). The TBA experiment showed that the presence of Mn/GAC (0.10) significantly promoted the generation of ·OH, which indicated that the catalyst accelerated the conversion of ozone into ·OH. To investigate how the interaction between ozone and catalyst affects the free radical generation mechanism, different amounts of dissolved ozone both with and without catalyst, as well as the results obtained from FTIR, XPS, EDS, and CV, were comprehensively analyzed.

Ozone and organic compounds can adsorb onto the surface of a catalyst and participate in catalytic ozonation reactions in heterogeneous catalytic oxidation systems (Bai et al. 2016; Zhu et al. 2017b). It has been reported that ozone molecules can form five-membered rings with oxygen-containing groups on the catalyst surface via hydrogen bonds and electrostatic forces. These molecules are then decomposed into free radicals (·OH) via electron transfer processes (Zhang and Ma 2008; Zhao et al. 2009a; Zhao et al. 2009b). The FTIR analysis (Fig. S3) indicated that the surface hydroxyl groups of GAC participated in the catalytic ozonation process. However, the addition of Fe or Mn oxides did not increase the hydroxyl group density on the surface, indicating that free radical generation was not dominated by the interactions between ozone and hydroxyl groups on the surface of the catalyst. XPS analysis showed that the Fe2+-Fe3+ and Mn2+-Mn3+-Mn4+ redox cycles were involved in the catalytic ozonation by Fe and/or Mn on GACs.

Previous studies showed that the recycling of Fe2+-Fe3+ and Mn2+-Mn3+-Mn4+ played key roles in promoting the generation of ·OH (Lv et al. 2010; Bai et al. 2016). The response mechanisms of these redox couples can be represented by the following equations (Xing et al. 2011; Wang and Bai 2017):

$$ \equiv {\mathrm{M}}^{\mathrm{n}+1}+{\mathrm{O}}_3\to \equiv {\mathrm{M}}^{\mathrm{n}+}+\cdotp \mathrm{OH}+{\mathrm{HO}}_2^{\cdotp }+{\mathrm{O}}_2 $$
(3)
$$ \equiv {\mathrm{M}}^{\mathrm{n}+}+{\mathrm{O}}_3+{\mathrm{H}}^{+}\to \equiv {\mathrm{M}}^{\mathrm{n}+1}+\cdotp \mathrm{OH}+{\mathrm{O}}_2 $$
(4)

Besides, the electron process, through the redox recycling of Fe2+-Fe3+ and Mn2+-Mn3+-Mn4+, is considered to be a significant process for catalytic ozonation (Lv et al. 2010). To further investigate the electron transfer process at the water-catalyst interface, the cyclic voltammetry behaviors of GAC and Fe and/or Mn/GAC electrodes were analyzed in a 3 wt% NaCl solution in an ozone atmosphere. Figure 6 shows that no significant redox current was observed on the GAC electrode, while both Fe- and/or Mn-modified electrodes showed obvious current signals. In addition, the order of the redox current was Mn/GAC (0.10) > Fe-Mn/GAC (0.10:0.10) > Fe/GAC (0.10), which is identical to the order of phenol mineralization efficiency. These results showed that both the strong interaction between catalyst and ozone and the enhanced interfacial electron transfer capability improved the catalytic activity of the modified catalysts. The above analyses show that the catalytic effects of Fe and/or Mn/GAC were mainly attributed to the recycling of Fe2+-Fe3+ and Mn2+-Mn3+-Mn4+, and changes in the concentration of dissolved ozone further demonstrated that the catalyst enhanced ozone’s ability to produce hydroxyl radicals.

Fig. 6
figure 6

Cyclic voltammetry scans of GAC, Fe/GAC (0.10), Fe-Mn/GAC (0.10:0.10), and Mn/GAC (0.10) electrodes in 3 wt% NaCl solution with ozone (22.0 mg L−1), Scan rate: 0.05 V s−1

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However, compared with the Mn/GAC (0.10)/O3 system, the Fe-Mn/GAC (0.10:0.10)/O3 system exhibited poorer TOC removal performance. This is inconsistent with a previous study which showed that combining FeOx and MnOx led to a synergistic effect which improved catalytic ozonation (Tang et al. 2016). Some literatures have showed that magnetite and MnO2 can form a galvanic corrosion coupling product, resulting in strong interactions on the catalyst surfaces, which cause Fe and Mn to simultaneously dissolve in neutral environments (Nakazawa et al. 2016; Saavedra et al. 2018). Fe and Mn can be released via galvanic interactions according to Eqs. (5) and (6):

$$ 3{\mathrm{Fe}}_3{\mathrm{O}}_4\to 4\ \left({\mathrm{Fe}}_2{\mathrm{O}}_3\ \mathrm{or}\ \mathrm{FeOOH}\right)+{\mathrm{Fe}}^{2+}+2{\mathrm{e}}^{-} $$
(5)
$$ {\mathrm{Mn}\mathrm{O}}_2+4{\mathrm{H}}^{+}+2{\mathrm{e}}^{-}\to {\mathrm{Mn}}^{2+}+2{\mathrm{H}}_2\mathrm{O} $$
(6)

XPS analysis indicated that magnetite (Fe3O4) and MnO2 were present on the surface of Fe-Mn/GAC, and carboxylic acids formed during the reaction process rapidly reduced the solution pH (Fig. S4). In addition, Nakazawa et al. (2016) showed that carbon materials could enhance galvanic corrosion by improving the conductivity between cathode and anode. Analysis of leaching metal ions (Table S4) showed that the concentrations of Fe and Mn ions in solution from the Fe-Mn/GAC (0.10:0.10)/O3 system were significantly higher compared with those of the Fe/GAC (0.10)/O3 and Mn/GAC (0.10)/O3 systems. This confirmed the above inference about the presence of galvanic corrosion on the Fe-Mn/GAC surface. Therefore, the poor catalytic performance of Fe-Mn/GAC (0.10:0.10) is attributed to the reduction in the number of active sites, which occurred mainly due to the galvanic corrosion of Fe3O4 and MnO2 which formed dissolved Fe2+ and Mn2+, respectively.

Therefore, in heterogeneous catalytic ozonation systems, the mineralization of organic compounds can be divided into two general reaction pathways: direct reactions with ozone and indirect reactions with free radicals generated from interactions between ozone and catalysts. Based on the above analyses, possible catalytic mechanisms of Fe and/or Mn/GAC are tentatively proposed and shown in Fig. 7.

Fig. 7
figure 7

Proposed mineralization mechanism of phenol in Fe or/and Mn/GAC catalytic ozonation systems

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The stability of catalysts

Since industrial applications require catalysts to remain active over a long period of time, the reusability of the prepared catalysts in the catalytic ozonation of phenol wastewater was examined. Figure 8a shows that no remarkable changes were observed in the catalytic performance of the prepared catalysts after the fifth run, and the TOC removal efficiencies was only decreased by 7.2% (Fe/GAC (0.10)) and 9.5% (Mn/GAC (0.10)). This indicated that both Fe/GAC (0.10) and Mn/GAC (0.10) were stable during continuous catalytic ozonation. However, the TOC removal efficiency decreased by 14.7% after the fifth run, which can be ascribed to fewer active catalytic sites due to galvanic corrosion on the surface of Fe-Mn/GAC (0.10:0.10). The leaching of Fe and Mn from Fe and/or Mn/GAC catalysts during catalytic ozonation was investigated using ICP-OES, as shown in Table S4. Based on the results, at a reaction time of 90 min, the concentrations of Fe and Mn in aqueous solution using the Fe/GAC (0.10)/O3 and Mn/GAC (0.10)/O3 systems were 0.06 mg L−1 and 0.07 mg L−1, respectively, which implies that Fe and Mn were stable on the catalysts. Fe (0.36 mg L−1) and Mn (0.45 mg L−1) were more likely to be released from the surface of Fe-Mn/GAC (0.10:0.10) compared with the Fe or Mn/GAC (0.10) catalysts, which is attributed to the galvanic corrosion of the Fe-Mn/GAC surface.

Fig. 8
figure 8

Catalytic performance of catalysts in successive ozonation runs (a) (Cphenol = 200.0 mg L−1; catalysts dosage = 1.0 g L−1; ozone concentration = 22.0 mg L−1; Each run time = 90 min); phenol, TOC, and COD removal and the evolution of NH4+-N and biodegradability index during catalytic ozonation (b); inset: color change of water sample before (a) and after (b) reaction (Mn/GAC (0.10) catalysts dosage = 1.0 g L−1; ozone concentration = 22.0 mg L−1)

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Application to secondary petrochemical effluent

To verify the performance of the catalytic ozonation system using actual wastewater, ozonation was used to treat secondary petrochemical effluent with Mn/GAC (0.10). As shown in Fig. 8b, when the phenol removal efficiency was 100%, the TOC decreased from 148.9 to 63.9 mg L−1 with a mineralization efficiency of 57.1%, and the COD decreased from 422.5 to 97.5 mg L−1 with a removal efficiency of 76.9%. The B/C increased from 0.011 to 0.045 after 90 min of catalytic ozonation, indicating the biochemical properties of the treated effluent were significantly improved. In addition, the NH4+-N concentration increased to 33.0 mg L−1 in the first 5 min, suggesting that some nitrogen–containing derivatives were produced during the catalytic ozonation of secondary petrochemical effluent. Afterwards, the NH4+-N concentration decreased to 0.8 mg L−1, which was attributed to indirect oxidation by ·OH and adsorption by the Mn/GAC (0.10) catalyst (Chen et al. 2017b; Kuang et al. 2018). In addition, it was found that the Mn/GAC (0.10) catalytic ozonation system significantly decolorized the secondary petrochemical effluent which was originally a transparent yellow liquid. The ecotoxicity evolution results in Fig. S5 show a slight toxicity increase in the first 5 min, which is possibly due to the generation of additional bio-inhibitory compounds during the degradation of phenol (Zhu et al. 2017b). However, the luminous intensity of Vibrio fischeri increased from ~ 450 to ~840 mV at 90 min of the catalytic ozonation process, suggesting that the process significantly decreased the toxicity of the secondary petrochemical effluent. In summary, catalytic ozonation based on Mn/GAC (0.10) notably improved the water quality index of secondary petrochemical effluent and efficiently treated secondary petrochemical effluent.

Engineering significance

The findings of this study have demonstrated that the Mn/GAC (0.10)/O3 catalyst is a stable and high-performance material for mineralizing phenol-containing wastewater. Due to this high efficiency, low cost, and simple operation of the catalyst, it can also be used for other refractory organic wastewaters. Due to the proper particle size of the catalyst (18–40 mesh), separation and recovery of the catalyst in heterogeneous catalytic ozonation was not an issue. The Mn/GAC (0.10)/O3 catalyst effectively treated secondary petrochemical effluent, and the results of this study improve the understanding of catalytic ozonation when used to treat secondary petrochemical effluent. In addition, this study also found that galvanic corrosion occurred in the Fe-Mn/GAC-based catalytic ozonation system in an acidic environment. This may provide a good starting point for further application and development of bimetallic catalysts in catalytic ozonation.

Conclusions

Fe and/or Mn/GAC catalysts were successfully prepared by precipitation-calcination and used for the treatment of phenol-containing wastewater. The catalytic performance order of the catalysts was Mn/GAC (0.10) > Fe-Mn/GAC (0.10:0.10) > Fe/GAC (0.10), and the Mn/GAC (0.10)/O3 system removed approximately 23.7% more TOC compared with the GAC/O3 system. Fe and/or Mn/GACs significantly enhanced the decomposition of O3 into ·OH, which played an important role in phenol mineralization. Some intermediate products of phenol degradation were identified, and a possible degradation pathway was proposed. The results of XPS, CV, and other experiments indicated that introducing Fe and Mn enhanced the interfacial electron transfer through Fe2+-Fe3+ and Mn2+-Mn3+-Mn4+ redox cycles, which increased the catalytic activity. Galvanic corrosion occurred between Fe3O4 and MnO2 on the surface of the Fe-Mn/GAC, which decreased its catalytic activity. Fe and Mn/GACs (0.10) still had high catalytic activities after 5 cycles, indicating that they were promising catalysts for the treatment of phenol-containing wastewater. The Mn/GAC/O3 system significantly improved the quality of actual secondary petrochemical effluent, including reducing the COD, TOC, NH4+-N, water color, and ecotoxicity, thus demonstrating that it had great potential in practical engineering applications. This study provides a new insight for the treatment of phenol-containing wastewater and secondary petrochemical effluent by catalytic ozonation.