The iron sludge generated from an electrocoagulation process was employed for the degradation of phenol through photo-Fenton process instead of iron salts. The morphology, functional groups and chemical composition of the iron sludge were investigated using scanning electron microscopy, Fourier infrared spectroscopy, and energy-dispersive X-ray spectroscopy. The effects of iron dose, H2O2 concentration, reaction time and initial concentration of phenol on photo-Fenton process performance were studied. The degradation efficiency of phenol reached 100% in the case of light compared to 68.6% in the case of dark reaction at pH of 3, iron dose of 150 mg/L and H2O2 concentration of 1.5 g/L. The degradation efficiencies of phenol were 100%, 71.3 and 51% at initial phenol concentrations of 50 mg/L, 100 mg/L and 150 mg/L, respectively. The expected cost for the treatment of one cubic meter of the contaminated wastewater was estimated to be 0.6224 $/m3.
Iron sludge was treated and employed as a catalyst in photo-Fenton process.
High degradation efficiency of phenol was attained.
The phenol degradation was improved under illumination.
Using iron sludge in photo-Fenton process reduced the total costs.
Phenol is frequently detected in ground and surface water through the discharge of effluents from food processing, resins and oil refinery industries (Gar Alalm et al. 2016). The existence of phenol in water streams even at low concentrations has attracted wide attention due to its adverse impact on human public health and aquatic life (Sánchez-Rodríguez et al. 2018). Moreover, the development of a new generation of bacteria that are capable of resisting a broad range of antibiotics and disinfectants may take place due to the presence of phenol in the aquatic environment (Jiang et al. 2019). The low biodegradability, as well as the toxicity of phenol, complicate its removal through conventional biological treatment processes (Samy et al. 2020a, 2020b). Moreover, physio-chemical processes are not effective to remove phenol, as they transform the non-biodegradable organics to sludge. The generation of high amounts of sludge requires post-treatment (Samy et al. 2019). Therefore, researchers have focused on the development of new techniques that have the potential to degrade bio-resistant pollutants to harmless intermediates. Recently, advanced oxidation processes (AOPs) have been employed for the removal of bio-recalcitrant contaminants through the generation of reactive species such as the hydroxyl radical (OH.) that are able to oxidize the bio-resistant pollutants to simpler by-products (Gar Alalm et al. 2018; Samy et al. 2020c). AOPs such as photocatalytic degradation (Shi et al. 2019; Samy et al. 2020d), Fenton reaction (Gümüş & Akbal 2016), photo-Fenton process (Wang et al. 2017) and ozonation (Wei et al. 2020) have been investigated for the degradation of phenol. Among AOPs, the photo-Fenton process has proved a high degradation performance for various refractory pollutants. The generation of hydroxyl radicals in the Fenton process happens via the reaction of ferrous ions (Fe2+) with hydrogen peroxide (H2O2) (Wang et al. 2019). The aforementioned reaction results in the oxidation of Fe2+ to ferric ions (Fe3+). In the photo-Fenton process, extra hydroxyl radicals can be generated, as the light reduces Fe3+ to Fe2+ to react again with H2O2 (Grassi et al. 2020). Moreover, the H2O2 can be decomposed by light to hydroxyl radicals (Wang et al. 2018). The additional hydroxyl radicals generated in the photo-Fenton process give this process its superiority over other traditional Fenton reactions.
Electrocoagulation (EC) is one of the previously mentioned physio-chemical processes that produce considerable quantities of sludge. The sludge generated from EC using iron electrodes contains iron hydroxides which can be employed as a catalyst in the photo-Fenton process. Therefore, in this study, the sludge taken from the EC process was utilized in the photo-Fenton reaction to overcome the problem of sludge related to the EC process and minimize the cost of the photo-Fenton process.
In this study, the iron sludge from the EC process was employed in the photo-Fenton process instead of iron salts. The morphology, chemical composition and functional groups of the iron sludge were studied using scanning electron microscopy (SEM), energy dispersive X-ray (EDX) spectroscopy and Fourier transform infrared spectroscopy (FTIR). The effects of reaction time, initial concentration of phenol, H2O2 concentration and the iron sludge dose were investigated. The degradation kinetics of phenol were studied. Moreover, estimation of the expected cost per cubic meter (m3) of the industrial wastewater containing phenol was carried out.
MATERIALS AND METHODS
Phenol (C6H5OH, 99.5%) was purchased from Sigma Aldrich. Methanol (CH3OH), hydrogen peroxide (H2O2), sodium hydroxide (NaOH) and hydrochloric acid (HCl) were purchased from Merck. All chemicals were used directly without any modifications.
Preparation of iron sludge
The sludge generated from the EC process was calcined at 500 °C for 4 h to remove all organics. Then, the sludge was cooled at room temperature for 24 h. After cooling, the sludge was ground and sieved with a 100 mesh sieve to ensure fine particles. The sludge was then vigorously washed with water to remove impurities. The final step was the drying of the washed sludge in an oven at 105 °C for 24 h.
All experiments were conducted in a photo-Fenton reactor. The reactor comprises a 250 mL Pyrex beaker containing 100 mL of the phenol solution at the desired concentration. A metal halide lamp (400 W, Philips) was set at 15 cm from the solution surface. The wavelength and photon flux of the lamp were 510 nm and 220 μW cm−2, respectively. Samples were withdrawn every 30 min and filtered before the analysis. The influences of various operating parameters were studied by changing the H2O2 concentration from 0 g/L to 1.5 g/L, initial phenol concentration from 50 mg/L to 150 mg/L and iron concentration from 0 mg/L to 200 mg/L.
The concentration of phenol was quantified using HPLC (Agilent 1200 series). All samples were filtered using a syringe filter with a porosity of 0.2 μm and then injected to the C-18 column. To detect the signal peak of phenol, a mixture of methanol and ultrapure water with ratios of 88 and 12%, respectively, was used as a mobile phase. The injection volume, temperature and flow rate were 25 μL, 40 °C and 0.9 mL/min. The detection wavelength was 270 nm.
The morphology of iron sludge was studied using scanning electron microscopy (SEM, Joel JEM 2100, Japan). Fourier transform infrared spectroscopy (FTIR, Nicolet IS 10, USA) was used to specify the functional groups of the iron sludge. The chemical composition of iron sludge was defined using energy dispersive X-ray (EDX, JEOL JSM-6510LV, Japan) spectroscopy.
RESULTS AND DISCUSSION
Characterization of iron sludge
The morphology of iron sludge particles is spherical as shown in Figure 1(a). The SEM image demonstrated the small size of iron sludge particles suggesting a high surface area. The well-crystalline surface, as well as the existence of some floccules from the EC process, are confirmed by SEM image. The EDX pattern in Figure 1(b) confirmed the existence of Fe, C, O and Ca. The presence of Ca2+ ions in the sludge components may be due to the precipitation of Ca2+ ions during the EC process by the reaction with carbonate anions. The presence of other elements (Si and Cu) may be due to the composition of the imaging grid (Bakre et al. 2016). Figure 1(c) shows the FTIR spectra of the iron sludge. The peaks at 463 cm−1 and 544 cm−1 are attributed to the Fe-O stretching vibration modes insuring the existence of iron hydroxides (Novoselova 2016). The bands at 1,630 cm−1 and 3,423 cm−1 are ascribed to the O-H stretching modes (Omwene & Kobya 2018). Moreover, the peaks in the region from 1,600 cm−1 to 1,650 cm−1 are assigned to H-O-H bending mode (Novoselova 2016). The bands observed at 2,927 cm−1, 2,966 cm−1 and 1,382 cm−1 are allocated to bending modes of C- H that come from the residual organics in the iron sludge (Samy et al. 2019).
Effect of illumination time, pH and light on the degradation efficiency of phenol
The degradation rates of phenol were high in the early stages of the reaction, then the degradation rates started to decline due to the consumption of hydroxyl radicals. The hydroxyl radicals were consumed in the removal of phenol and its intermediates. During the degradation of phenol, aromatic intermediates such as benzoquinone, catechol (benzene-1,2-diol), hydroquinone (1,4-Dihydroxybenzene) can be generated (Gar Alalm et al. 2016). Then, the continuous oxidation of the generated by-products converts the intermediates to simpler by-products such as maleic, oxalic, acetic, and glycol acids and finally transformed to CO2 and H2O (Gar Alalm et al. 2016). The oxidation of these intermediates via hydroxyl radicals reduced the hydroxyl radicals available for the degradation of phenol and consequently, the degradation rates of phenol declined with time.
The photo-Fenton process is greatly affected by pH. Previous studies confirmed the high degradation performance of the photo-Fenton process in acidic conditions, so the preliminary experiments were performed at pH values of 2, 3, 4 and 5 to specify the optimum pH value (Radwan et al. 2018). At pH higher than 3, the degradation efficiency declined due to the formation of Fe(OH)3 resulting in the reduction of Fe+3 ions (Shokri 2018). Moreover, the continuous generation of Fe+2 can be inhibited, leading to lower photodegradation efficiency of phenol (Rosales et al. 2012). At a pH value of 2, H2O2 employs as a hydroxyl radicals deactivator (Radwan et al. 2018). Furthermore, oxonium ions probably exist in the case of low values of pH resulting in the increase of H2O2 stability (Xu et al. 2009). The increase of H2O2 stability reduces the probability of the reaction between H2O2 and Fe2+. The results showed that high degradation efficiency of phenol was achieved at a pH of 3 which is in accordance with our previous study (Radwan et al. 2018). Therefore, all the following experiments were conducted at pH of 3.
Effect of H2O2 concentration
The increase of H2O2 concentration from 0 g/L to 1.5 g/L, as shown in Figure 3, enhanced the degradation efficiency from 12.6% to 100% at pH of 3, iron dosage of 1.5 g/L and initial phenol concentration of 50 mg/L due to the rise of hydroxyl radicals generated from the reaction between Fe+2 and H2O2 as well as the hydroxyl radicals released from the decomposition of H2O2 by light. It was noticed that the degradation of phenol using 1 g/L of H2O2 was 94.5% and the degradation efficiency was 100% at an H2O2 concentration of 1.5 g/L. The increase of H2O2 concentration above 1 g/L did not significantly improve the degradation efficiency due to the formation of hydroperoxyl radicals (Tolba et al. 2020). The oxidation ability of hydroperoxyl radicals is lower than hydroxyl radicals (Zamora et al. 2006). Moreover, higher concentrations of hydrogen peroxide may employ as a scavenger of hydroxyl radicals (Xie et al. 2016). Furthermore, the reaction of hydrogen peroxide with hydroxyl radicals instead of phenol may take place at high concentrations of H2O2 (Elmolla & Chaudhuri 2010).
Effect of iron sludge dose
The rise of iron dose from 0 mg/L to 150 mg/L improved the degradation efficiency from 59.6% to 100% due to the increase of hydroxyl radicals generated, as shown in Figure 4. The degradation efficiency of phenol was 100% after 240 min at an iron dose of 125 mg/L, pH of 3 and initial phenol concentration of 50 mg/L, whereas the removal efficiency of phenol was 100% after 180 min at an iron dose of 150 mg/L, pH of 3 and initial phenol concentration of 50 mg/L. However, the raising of iron dose above 150 mg/L may have a reverse impact on the degradation efficiency of phenol. The decline of degradation efficiency may be due to the consumption of iron in the coagulation of fine particles (Adel et al. 2020). Moreover, higher doses of iron lead to the high turbidity in the solution hindering the light to reduce Fe3+ and decompose H2O2 (Shokri 2018).
Effect of the initial concentration of phenol
The effect of the initial concentration of phenol on the degradation efficiency of phenol is shown in Figure 5(a). The removal efficiencies of phenol were 100%, 71.3 and 51% at initial phenol concentrations of 50 mg/L, 100 mg/L and 150 mg/L, respectively at pH of 3, iron dose of 150 mg/L and H2O2 concentration of 1.5 g/L. The increase of initial phenol concentration decreased the degradation efficiency of phenol, as the number of hydroxyl radicals generated remained constant regardless of the initial concentration of phenol (Saber et al. 2014). Therefore, the reaction time should be extended in the case of higher concentrations of phenol to achieve the desired removal.
The TOC removal efficiency of 50 mg/L of phenol was 68.6%, as shown in Figure 5(b). The TOC mineralization rate was high at the beginning of the reaction, then started to decrease due to the rise of generated intermediates. The previously mentioned by-products increase the TOC concentration in the solution. However, the TOC values continued to decrease, as the degradation rate remained higher than the by-products generation rate. To achieve a high mineralization percentage of TOC, the reaction time should be extended.
Degradation kinetics of phenol
The degradation rates were 0.0104, 0.0051 and 0.0031 at initial phenol concentrations of 50 mg/L, 100 mg/L and 150 mg/L, respectively. The R2 values were 0.92, 0.94 and 0.97 in the case of initial phenol concentrations of 50 mg/L, 100 mg/L and 150 mg/L, respectively.
The estimation of photo-Fenton process cost
It is estimated that the cost of chemicals and energy consumption cost are 0.2 $/m3 and 0.3 $/m3, respectively. Therefore, the operating cost including maintenance cost is 0.5024 $/m3 and the total cost is .6224 $/m3.
High degradation efficiency of phenol was achieved using iron sludge from the EC process.
The characterization of iron sludge confirmed the small size of sludge particles and the existence of iron hydroxides.
The existence of light plays a major role in the enhancement of degradation efficiency of phenol.
The increase of H2O2 concentration above 1 g/L did not greatly improve the degradation efficiency.
The raising of the iron dose above 150 mg/L decreased the degradation efficiency of phenol.
The degradation efficiency of phenol went down by increasing the initial concentration of phenol.
The degradation of phenol followed the pseudo-first order model.
The use of sludge from the EC process instead of iron salts in the photo-Fenton process reduced the cost of the treatment of one cubic meter of the contaminated wastewater and reuse of iron sludge decreases the environmental burden.
The experimental work was supported by the Science and Technology Development Fund (STDF) (project No 26279). The second author is supported by a postdoctoral fellowship granted by the Japan Society for the Promotion of Science (JSPS) (Grant No: 19F19349).
CONFLICT OF INTEREST
The authors declare that there are no conflicts of interest in this work and its publication.
DATA AVAILABILITY STATEMENT
All relevant data are included in the paper or its Supplementary Information.