*Article* **CoFe2O4 Nanomaterials: E**ff**ect of Annealing Temperature on Characterization, Magnetic, Photocatalytic, and Photo-Fenton Properties**

**Nguyen Thi To Loan 1,\*, Nguyen Thi Hien Lan 1, Nguyen Thi Thuy Hang 2, Nguyen Quang Hai 3, Duong Thi Tu Anh 1, Vu Thi Hau 1, Lam Van Tan 4,5 and Thuan Van Tran 4,6,\***


Received: 23 October 2019; Accepted: 21 November 2019; Published: 28 November 2019

**Abstract:** In this research, structural, magnetic properties and photocatalytic activity of cobalt ferrite spinel (CoFe2O4) nanoparticles were studied. The samples were characterized by X-ray powder diffraction (XRD), energy dispersive X-ray (EDX), scanning electron microscopy (SEM), transmission electronic microscopy (TEM), Brunauer–Emmett–Teller (BET), Fourier transform infrared spectroscopy (FTIR), and UV-visible diffused reflectance spectroscopy (DRS) analysis. The XRD analysis revealed the formation of the single-phase CoFe2O4 with a cubic structure that is annealed at 500–700 ◦C in 3 h. The optical band gap energy for CoFe2O4 was determined to be in the range of 1.57–2.03 eV. The effect on the magnetic properties of cobalt ferrites was analyzed by using a vibrating sample magnetometer (VSM). The particle size and the saturation magnetization of cobalt ferrite nanoparticles increased with increasing annealing temperature. The photocatalytic activity of CoFe2O4 nanoparticles was investigated by using rhodamine B dye under visible light. The decomposition of rhodamine B reached 90.6% after 270 min lighting with the presence of H2O2 and CF500 sample.

**Keywords:** cobalt ferrite; magnetic properties; solution combustion method; rhodamine B; photocatalytic activity

#### **1. Introduction**

Among many ferrites, cobalt ferrite magnetic nanoparticles are attracting much attention because of their high coercivity, magnetocrystalline anisotropy, moderate saturation magnetization, chemical stability, wear resistance, electrical insulation, and structure [1]. Structurally, in the inverse spinel of the ferrite, tetrahedral sites are generally occupied by Fe3<sup>+</sup> ions, whereas octahedral sites (B-sites) are inhabited by Co2<sup>+</sup> and Fe3<sup>+</sup> ions [2]. To alter structure and magnetic properties of ferrite nanoparticles, it is necessary to modify their composition and microstructures via different preparation routes [2]. CoFe2O4 nanoparticles were previously prepared by a wide array of synthesis routines, such as sol-gel [3,4], hydrothermal method [5], chemical co-precipitation [6,7], solvothermal [8], solid-state method [9], and solution combustion [10–13]. For each synthesis method, it was found that the

annealed temperature played a key role in determining the structure and properties of the obtained product. In recent years, photocatalytic oxidation of various dyes using ferrites has drawn a great deal of attention, opening new trends in the environmental remediation [14–16].

Currently, contamination of the water environment has been alarming due to the widespread use of organic compounds in manufacturing processes and the rapid development of dyeing industries. These dye compounds existing in water pose a direct threat to public health and to animal and aquatic life due to their toxicity, endocrine-disrupting capability, and mutagenic or potentially carcinogenic properties [17]. One of the most effective methods of solving this problem is advanced oxidation processes (AOPs). Among different AOPs, the photo-Fenton-like reaction was widely studied because it can produce more oxidative species such as hydroxyl radicals (•OH) to accelerate the reaction [18]. This method is based on the use of semiconductors and oxidant and light sources to perform the decomposition of organic matter. The advantages of the photo-Fenton processes consist of environmental friendliness and the ability to decompose completely organic pollutants into non-toxic inorganic substances, these being CO2 and H2O. Spinel ferrite materials have received wide application as photocatalysts due to their structural composition and thermal and chemical stability toward various reaction conditions [18]. The photo-Fenton reagent using ferrite can be easily recovered from the solution by an external magnetic field and is available for reutilization.

Abul Kalam et al. reported the photocatalytic activity of cobalt ferrite magnetic nanoparticles for degradation of methylene blue with H2O2 under visible light irradiation, and achieved very good performance [15]. P. Annie Vinosha et al. synthesized NiFe2O4 by co-precipitation technique. The photocatalytic application for the synthesized sample was studied for the degradation of methylene blue dye. In the presence of H2O2 and ferrites, methylene blue degradation efficiency reached ~30% in the dark but degradation improved to ~99% in the irradiation light [19]. MgFe2O4 nanoparticles synthesized by a solution combustion method exhibited a high ability for Photo-Fenton-like degradation of methylene blue [20]. In the photo-Fenton processes, hydrogen peroxide is used commonly as an oxidant. Hydrogen peroxide is safe and easy to handle and poses no lasting environmental threat because it readily decomposes to water and oxygen [17].

One of the major pollutants discharged from various industries is dyes [21–23]. Previous reports have revealed that thousands of new dyes have been synthesized and commercialized, with the total amount of approximately one million tons being consumed throughout the world [24,25]. Ever-increasing utilization and direct discharge without treatment of colored effluents have been considered as a problematic obstacle, affecting the photosynthesis of aquatic lives because of the reduction of the ability of light penetration [26]. Amid the most emergent synthetic dyes, rhodamine B (RhB) is a virtually hazardous and non-biodegradable dye [27]. In chemical essence, this compound is categorized as a cationic and soluble dye, in accordance with the existence of highly stable tertiary amine and carboxylic groups in its molecular structure. It is thus found to have a profound impact on living creatures as well as ambient environment via a range of approaches on direct or indirect exposure [28]. With such harmful and dangerous properties to many organisms, effective removal of dyes from wastewater is essential, but currently remains a challenge [29].

To eliminate the RhB contamination in wastewater, the adoption manifold feasible methods has been suggested, involving adsorption [30], electro-Fenton process [31], and microfiltration membrane [32]. For instance, Tawfik et al. synthesized the nano-sized polyamide-grafted carbon microspheres via interfacial polymerization, exhibiting a promising adsorption performance towards RhB at 19.9 mg/g [33]. Recently, Mustafa et al. have successfully attained a kind of eco-friendly activated carbon-based modified nanocomposite that combined carbon with bimetallic Fe and Ce nanoparticles [34]. Regardless of giving promising results in high surface area (~423 m2·g−1) and adsorption capacity (324.6 m2·g−1) towards RhB, the complicated preparation procedure limits applications of the synthesized nanocomposite. On the other hand, although the introduction of minerals as adsorbents for treating RhB have been developed, such approaches seem to confront many obstacles relating to material stability, recyclability, fabrication cost, and the reliance on post-treatment

separation [35–37]. With high stability, magnetism, and photocatalysis, CoFe2O4 nanoparticles are expected to deal with RhB pollution in water efficiently.

In this present report, the spinel cobalt ferrites nanoparticles are characterized for their structural, morphological, optical, and magnetic properties using various methods. In addition, the photocatalytic activity of samples was investigated by the degradation process of rhodamine B.

#### **2. Materials and Methods**

#### *2.1. Materials*

Cobalt nitrate hexahydrate (Co(NO3)2·6H2O), iron nitrate nonahydrate (Fe(NO3)3·9H2O), urea (CH4N2O), rhodamine B (C28H31ClN2O3), and hydrogen peroxide were obtained from Merck and used as received, without further purification.

#### *2.2. Synthesis of CoFe2O4 Nanoparticles*

The synthesis of cobalt ferrite was performed via solution combustion method using urea as fuel [11]. The process commenced with the dissolution of urea in the water, followed by the addition of Co(NO3)2·6H2O and Fe(NO3)3·9H2O at an appropriate amount and mole ratio under vigorous stirring. The resultant mixed solution was stirred further to afford a gel, which was then subjected to heating at 70 ◦C for 12 h in an oven. The obtained powder product was calcined at four different temperatures ranging from 500 to 800 ◦C for 3 h with a heat rate of 5 ◦C min<sup>−</sup>1, and the subsequent products were labeled as CF500, CF600, CF700, and CF800 respectively.

#### *2.3. Characterizations*

The first characterization involved determining the crystallite size, r, of spinel using Scherrer's formula as follows:

$$r = \frac{k\lambda}{\beta \cos \theta} \tag{1}$$

where λ, *k*, β*,* and θ are the X-ray wavelength, Scherrer's constant (*k* = 0.89), the full width at half maximum observed in radians, and the angle of diffraction of the (311) peak with the highest intensity, respectively. θ and β were obtained via X-ray diffraction (XRD) results using a D8 Advance diffractometer (Brucker, Madison, WI, USA) instrument with CuK<sup>α</sup> radiation (λ = 1.5406 Å) in a 2θ angle ranging from 20 to 70◦ with a step of 0.03◦.

The second characterization regarding morphology of the particles was obtained via scanning electron microscope (SEM, Hitachi S-4800, Tokyo, Japan) and transmission electron microscopy (TEM, JEOL-JEM-1010, Tokyo, Japan). The composition of the samples was analyzed by energy dispersive X-ray spectroscopy (EDX, JEOL JED 2300 Analysis Station, Tokyo, Japan). N2-sorption investigation was performed to obtain the Bet-specific surface area of the product. A surface analyzer instrument (a Quantachrome Nova 2200, Boynton Beach, FL, USA) was employed to obtain the isotherm at 77 K. The specific surface area was calculated via the Brunauer–Emmet–Teller (BET) method. The spinel structure was affirmed by Fourier transform infrared spectroscopy (FTIR Affinity-1S, Shimadzu, Tokyo, Japan). A UV-VIS absorption spectrometer (U-4100, Hitachi, Japan) was employed to obtain the optical absorption spectra. To elaborate magnetic properties of samples, a vibrating sample magnetometer (VSM, Ha Noi, Vietnam) operating at room temperature was utilized.

#### *2.4. Photocatalytic Degradation of Rhodamine B*

Multiple ferrite samples synthesized at different annealing temperatures were examined for photodegradation performance against Rhodamine B (RhB) aqueous solution. The irradiation source was 30 W LED lamps (Philips). The experiment commenced with the addition of 100 mg of ferrite sample into 100 mL of 10 mg. L−<sup>1</sup> RhB aqueous solution. Following that, stirring was carried out in the dark for 30 min to allow the solution to reach the adsorption–desorption equilibrium state, followed by the addition of 1.5 mL of 30% H2O2. Consequently, visible light irradiation started and the reaction took place under stirring. After specific periods, 5 mL of aliquots were removed and subjected to centrifugation for particle separation. RhB concentration was determined using an ultraviolet-visible spectrophotometer (UV-1700 Shimadzu, Tokyo, Japan) at 553 nm. The degradation efficiency of the ferrite against RhB (H) was calculated as follows:

$$\mathrm{pH} = \frac{\mathrm{C\_o} - \mathrm{C\_t}}{\mathrm{C\_o}} \times 100\% \tag{2}$$

where C is the concentration of RhB. The subscript 0 and t denote equilibrium state and time (t) after irradiation, respectively.

#### **3. Results and Discussion**

#### *3.1. Structural Analysis*

Different XRD patterns of cobalt ferrites corresponding to different annealing temperatures are shown in Figure 1. The reflection peaks corresponded to the characteristic spacing between (220), (311), (400), (511), and (440) plans of a cubic spinel structure, providing clear evidence of the formation of cobalt ferrite (JCPDS number 22–1086) [5]. α-Fe2O3 peaks corresponding to secondary impurities were observed for the sample annealed at 800 ◦C, which was possibly caused by sample decomposition [38]. Table 1 lists the average crystallite size (r), calculated using (1), and the lattice parameter (a), obtained using the formula a<sup>2</sup> = d2/(h<sup>2</sup> + k2 + l 2) with inputs obtained from X-ray diffraction data. The mean crystallite size ranged from 9 to 29 nm and increased with elevated annealing temperature [29]. The lattice parameter for the samples of cobalt ferrites nanoparticles varied from 8.3347 to 8.3745 Å.

**Figure 1.** X-ray diffractions of CoFe2O4 nanoparticles annealed at 500–800 ◦C.


**Table 1.** Average crystallite size (r), lattice parameter (a), unit cell volume (V), wave number (ν<sup>1</sup> and ν2) for the tetrahedral and octahedral site, respectively, and Brunauer–Emmet–Teller (BET) surface area of all the CoFe2O4 samples.

FTIR spectra of CoFe2O4 nanoparticles annealed at different temperatures are displayed in Figure 2. The cobalt ferrite samples exhibited two vibration bands at wave number 522–532 cm−<sup>1</sup> (ν1) and at 408–412 cm−<sup>1</sup> (ν2), corresponding to the stretching vibration of the M-O bond in tetrahedral and octahedral sites (Table 1). A. Kalam et al. [15] observed that the vibration mode between tetrahedral metal ion and oxygen complex gives rise to the high-frequency band in the range of 597–615 cm−1, whereas stretching vibration between octahedral metal ion and oxygen complex gives rise to the weak frequency band in the range of 412–400 cm−<sup>1</sup> in the case of cobalt ferrite, which is an inverse spinel ferrite.

The FTIR results confirmed that the samples had a spinel structure of CoFe2O4, which was revealed by the XRD results.

**Figure 2.** Fourier transform infrared (FTIR) spectrum of CF500–CF800 samples.

Figure 3 shows the SEM micrographs of CoFe2O4 nanoparticles annealed at different temperatures. The SEM images indicated that particles were agglomerated and spherical. In addition, the crystalline size of the samples seemed to increase proportionally with annealing temperature, which was consistent with the results of XRD analysis. The agglomeration could be explained by the interaction between magnetic particles that occurred during annealing under high temperatures. In addition, it was previously found that higher annealing temperature inevitably causes moderate agglomeration [2].

**Figure 3.** Scanning electron microscopy (SEM) of CoFe2O4 nanoparticles: (**a**) CF500, (**b**) CF600, (**c**) CF700, (**d**) CF800.

TEM images (Figure 4a) of the CoFe2O4 annealed at 500 ◦C revealed that the particle size was approximately 20 nm. Compositional determination was performed by energy dispersive X-ray spectroscopy (EDX) analysis, showing peaks corresponding to Co, Fe, and O elements of the CF500 sample (Figure 4b).

**Figure 4.** Transmission electronic microscopy (TEM) (**a**) and energy dispersive X-ray (EDX) (**b**) of the CF500 sample.

Isotherm of N2 adsorption–desorption of products annealed at 500–800 ◦C are displayed in Figure 5. The decrease in the cobalt ferrites surface area was observed with the increase in annealing temperature (Table 1). The CoFe2O4 nanoparticles synthesized at 500 ◦C achieved the highest specific surface area of 12.69 m2·g<sup>−</sup>1.

Thus, annealing temperature affected the particle size, morphology, and surface area of cobalt ferrites.

**Figure 5.** N2 adsorption–desorption isotherm of CF500, CF600, CF700, and CF800.

Kubelka–Munk model was employed to calculate the band gaps (Eg) of CoFe2O4 nanoparticles using the absorption coefficient (α) calculated from diffuse reflectance spectra (DRS), as follows [39,40]:

$$\mathbf{F(R)} = \mathbf{x} = \frac{\left(1 - \mathbf{R}\right)^2}{2\mathbf{R}} \tag{3}$$

where F(R) is the Kubelka–Munk function, α is the absorption coefficient, and R is the reflectance. The band gap energy (Eg) of CoFe2O4 nanoparticles can be calculated by the following equation:

$$\mathbf{A}\mathbf{a}\mathbf{h}\mathbf{v} = \mathbf{A}(\mathbf{h}\mathbf{v} - \mathbf{E}\_{\mathbf{g}})^{\mathbf{n}} \tag{4}$$

where hν, α, A, and n represent energy of the photon, the absorption coefficient, the material parameter, and the transition parameter, respectively. n = 2 represents indirect transitions [41]. To determine the optical band gap energy (Eg), (αhν) <sup>2</sup> was plotted against photon energy (hν) to produce multiple Wood–Tauc plots (Figure 6). The band gap values of CF500, CF600, CF700, and CF800 samples were found to be 1.57, 1.66, 1.90, and 2.03 eV, respectively. This indicated that the annealing affected the optical band gap energy of CoFe2O4 nanoparticles. The optical band gap energy value increased with increasing temperature in annealed samples.

**Figure 6.** Wood–Tauc plots for CoFe2O4 nanoparticles: (**a**) CF500, (**b**) CF600, (**c**) CF700, (**d**) CF800.

#### *3.2. Magnetic Properties*

The magnetic properties of ferrites could be largely determined by various elements such as density, grain size, anisotropy, and A–B exchange interactions [1]. In this investigation, obtained ferrite products annealed at different temperatures were subjected to M–H hysteresis measurements carried out at room temperature. Various magnetic properties including saturation magnetization (Ms), coercivity (Hc), and remanent magnetization (Mr) are listed in Figure 7 and Table 2. It was observed that Ms value showed a positive correlation with particle size. This was in line with Kumar et al., suggesting that increased particle size could lead to improved magnetization [2].


**Table 2.** Magnetic parameter of the CoFe2O4 nanoparticles.

**Figure 7.** Hysteresis loop of CoFe2O4 nanoparticles: (**a** CF500, (**b**) CF600, (**c**) CF700, (**d**) CF800.

#### *3.3. Photochemical Activities*

Figure 8 represented the photocatalytic degradation of RhB versus irradiation time in different conditions. Generally, absorption peaks observed at 501 and 553 nm could be attributable to the absorption fully de-ethylated and tetraethylated rhodamine B molecules, respectively [41]. The role of oxidant and catalyst was further elaborated by performing reactions under typical conditions that were neither H2O2 nor CoFe2O4. Firstly, RhB degradation efficiency reached a marginal percentage, at just 10.2%, under visible light after 270 min in the sole presence of H2O2. This implied that RhB dye could be hardly photodegraded in the absence of CoFe2O4 catalyst. In the presence of CoFe2O4 sample without oxidation reagent H2O2, RhB degradation efficiency achieved 9.4% (in the dark) and 32.5% (under light irradiation) after 270 min. In the absence of light, 21.3% of RhB was degraded while using both CoFe2O4 and H2O2. Abul et al. also observed similar results by using CoFe2O4 as a catalyst for the degradation of methylene blue in liquid under air atmosphere [15].

Grasping this improvement, we speculated about an even better photodegradation efficiency when CoFe2O4 and H2O2 were combined for the next photoreaction. Photocatalyzed degradation efficiency (%) towards RhB under visible light irradiation and in the presence of CoFe2O4 and H2O2 against the interval irradiation time is indicated in the UV-visible spectra in Figure 9. Indeed, all the samples of CoFe2O4 (CF) nanoparticles offered an enhancement in catalytic performance, but the percentages of RhB decomposition varied according to the distinct annealing temperature (500, 600, 700, and 800 ◦C). The photocatalytic degradation efficiency of RhB was evaluated at about 90.6%, 67.6%, 51.6%, and 42.8% after 270 min of lighting in the presence of H2O2 and CF500, CF600, CF700, and CF800, respectively.

**Figure 8.** The plots of (*C*/*Co*) ×100 versus irradiation time (*t*) in different conditions: (**1**) H2O2, (**2**) CF500 + dark, (**3**) CF500 + light, (**4**) CF500 + H2O2 + dark.

**Figure 9.** UV-visible spectra of RhB degraded by (**a**) H2O2 + CF500, (**b**) H2O2 + CF600, (**c**) H2O2 + CF700, (**d**) H2O2 + CF800 after 270 min of lighting.

The pollutant degradation rate could be substantially increased due to highly oxidative hydroxyl ( •OH) radicals created by the heterogeneous photo-Fenton process, in the presence of spinel ferrite as a heterogeneous catalyst, using H2O2 as oxidant under light irradiation [15,42]. The mechanism of the heterogeneous photo-Fenton reaction was shown according to the following equations:

$$\text{Fe}^{3+} + \text{H}\_2\text{O}\_2 + \text{hv} \rightarrow \text{Fe}^{2+} + \text{HOO}^\bullet + \text{H}^+$$

$$\text{Fe}^{2+} + \text{H}\_2\text{O}\_2 + \text{hv} \rightarrow \text{Fe}^{3+} + \text{"OH} + \text{OH}^-$$

where Fe3<sup>+</sup> and Fe2<sup>+</sup> represent the iron species on the surface of a heterogeneous catalyst.

Due to Fe(II, III) cycling, the stability of the ferrite system is maintained during the degradation process and the active species are generated continuously [15].

The peak photocatalytic degradation efficiency for CF500 (90.6%) was due to the effective crystallite size (9 nm), separation, and prevention of electron-hole pair (e−/h+) recombination [43]. Photocatalytic performance of CF500 with the largest degradation rate could be interpreted via plausible hypotheses based on the following foundations:


Figure 10 depicted the plots with different pseudo first order rates, which were obtained by fitting the following equation:

$$\ln\left(\frac{\mathcal{C}\_o}{\mathcal{C}\_t}\right) = kt$$

where C is the RhB concentration. The subscripts 0 and t denote initial state and time t after irradiation. k is the pseudo first order rate kinetics.

With the higher the coefficient of determination (*R*<sup>2</sup> > 0.9), the proposed model was highly compatible [45,46]. The estimated parameters, including pseudo first order rate constant k values and *R*<sup>2</sup> values, are shown in Table 3. The first order rate constant for CF500 was 0.839 <sup>×</sup> 10−<sup>2</sup> min−1, and it was 3.9 times faster than that of CF800. High *R*<sup>2</sup> values also confirmed the adherence of the photocatalytic degradation of RhB to the first order kinetics.

**Figure 10.** The plots of ln(*Co*/*Ct*) versus irradiation time (*t*) in the presence of H2O2 and CoFe2O4 nanoparticles: (**1**) CF500, (**2**) CF600, (**3**) CF700, (**4**) CF800.


**Table 3.** Pseudo first order rate constant (k) for the photocatalytic degradation of RhB using CoFe2O4 nanoparticles.

#### **4. Conclusions**

CoFe2O4 spinel nanoparticles were successfully synthesized via solution combustion method using urea as a fuel. The effect of annealing temperature on phase composition, morphology, optical property, and magnetic properties of CoFe2O4 materials was studied. The crystallite size calculated by the Scherer formula increased from 9 to 29 nm with an increase in the annealing temperature. By elevating the annealing temperature, it was found that the band gap energy value and saturation magnetization of CoFe2O4 spinel were also accordingly increased. The photocatalytic degradation against RhB dye of CoFe2O4 spinel decreased with increasing annealed temperature. Among all the cobalt ferrite samples, CF500 exhibited an enhanced degradation efficiency of 90.6% at a visible light exposure time of 270 min. The first-order rate constant for CF500 was 0.839 <sup>×</sup> 10−<sup>2</sup> min−<sup>1</sup> and was 3.9 times faster than CF800. The photocatalytic degradation of RhB dye followed first order kinetics. The current results suggest a possible application of cobalt ferrites nanoparticles in treatment of dye-contaminated water.

**Author Contributions:** Investigation, N.T.H.L., N.T.T.H., N.Q.H., D.T.T.A., V.T.H., L.V.T. and T.V.T.; Writing—original draft, N.T.T.L.

**Funding:** This research received no external funding.

**Conflicts of Interest:** The authors declare no conflict of interest.

#### **References**


© 2019 by the authors. Licensee MDPI, Basel, Switzerland. This article is an open access article distributed under the terms and conditions of the Creative Commons Attribution (CC BY) license (http://creativecommons.org/licenses/by/4.0/).

#### *Article* **Degradation of 2,4-Dichlorophenol by Ethylenediamine-***N***,***N-*  **disuccinic Acid-Modified Photo-Fenton System: Effects of Chemical Compounds Present in Natural Waters**

**Wenyu Huang 1,2,3, Ying Huang 1, Shuangfei Wang 2,3, Hongfei Lin <sup>3</sup> and Gilles Mailhot 4,\***


**Abstract:** This paper describes a study of the treatment of 2,4-dichlorophenol (2,4-DCP) with an ethylenediamine-*N*,*N* -disuccinic-acid (EDDS)-modified photo-Fenton system in ultrapure water and different natural waters. The results showed that the EDDS-modified photo-Fenton system is adequate for 2,4-DCP degradation. Compared with a medium containing a single organic pollutant, the removal of pollutants in a more complex medium consisting of two organic compounds is slower by around 25 to 50% as a function of the organic pollutant. Moreover, 2,4-DCP can be further effectively degraded in the presence of organic materials and various inorganic ions. However, the photodegradation of 2,4-DCP in different natural waters, including natural lake water, effluent from domestic sewage treatment plants, and secondary effluent from pulp and paper mill wastewaters, is inhibited. Chemical compounds present in natural waters have different influences on the degradation of 2,4-DCP by adopting the EDDS-modified photo-Fenton system. In any case, the results obtained in this work show that the EDDS-modified photo-Fenton system can effectively degrade pollutants in a natural water body, which makes it a promising technology for treating pollutants in natural water bodies.

**Keywords:** photo-Fenton; EDDS; 2,4-DCP; organic matter; inorganic ions; natural water bodies

#### **1. Introduction**

Advanced oxidation processes (AOPs) can effectively oxidize organic pollutants in water using active free radicals [1], which have been adopted to degrade pollutants from many types of wastewaters, such as tannery wastewaters and pharmaceutical wastewaters [2,3]. Fenton technology is one of the most simple and effective AOPs to degrade pollutants [4]. However, the Fenton reaction consumes a large number of chemical reagents and produces toxic by-products, resulting in secondary contamination [5]. In order to overcome the shortcomings of Fenton processes, different attempts have been carried out. First of all, irradiation has been introduced in the Fenton reaction, and the Fe(III)/H2O2 mixture (Fenton-like reagent) can absorb photons of wavelengths up to 550 nm [6,7]. The efficiency of the photo-Fenton process has been proven by many previous studies, most of which were carried out at an acidic pH value [8–10]. Additionally, the introduction of chelating agents, especially organic carboxylic acid into the Fenton reaction, proved to be an effective method for broadening the applicable pH values. Several different organic carboxylic acids, such as citric acid, oxalic acid, NTA, and EDTA, were used to modify the Fenton process and proved to be efficient [11–14]. Ethylenediamine-*N*,*N* -disuccinic-acid (EDDS), a biodegradable isomer of EDTA, was used as a chelating agent in homogeneous

**Citation:** Huang, W.; Huang, Y.; Wang, S.; Lin, H.; Mailhot, G. Degradation of 2,4-Dichlorophenol by Ethylenediamine-*N*,*N* -disuccinic Acid-Modified Photo-Fenton System: Effects of Chemical Compounds Present in Natural Waters. *Processes* **2021**, *9*, 29. https://dx.doi.org/ 10.3390/pr9010029

Received: 24 November 2020 Accepted: 23 December 2020 Published: 25 December 2020

**Publisher's Note:** MDPI stays neutral with regard to jurisdictional claims in published maps and institutional affiliations.

**Copyright:** © 2020 by the authors. Licensee MDPI, Basel, Switzerland. This article is an open access article distributed under the terms and conditions of the Creative Commons Attribution (CC BY) license (https://creativecommons.org/ licenses/by/4.0/).

and heterogeneous Fenton and photo-Fenton systems in our previous studies [15–18]. Therefore, the EDDS-modified photo-Fenton reaction is proven to be a promising approach to treat refractory pollutants.

However, most laboratory studies on pollutant removal using the EDDS-modified photo-Fenton system are currently conducted using deionized water, which is far from the complex chemical composition of natural water, in which the inorganic ions and dissolved organic compounds in water can significantly influence pollutant removal. In previous research, it was found that the water matrix could significantly influence the efficiency and mechanism of AOPs processes, especially the Fenton process. The presence of common inorganic ions had no substantial effect on herbicide removal when using the photo-Fenton system, but the H2O2 (oxidant) consumption of this reaction was higher than that of the same reaction where inorganic ions were absent [19]. In the process of the photodegradation of 2,4-D by the photo-Fenton reaction, carbonate has little adverse effects, fluoride has a positive effect, and phosphate has an inhibitory effect [20]. The effect of typical inorganic water constituents (carbonates and chloride ions) and organic matter was also investigated and found to be different in the UVA-UVB activation of hydrogen peroxide and persulfate for advanced oxidation processes [21]. Therefore, it is necessary to study the influence of the natural water matrix on pollutant degradation efficiency in the EDDS-modified photo-Fenton system, including not only single inorganic ions or organic compounds but also natural water where various chemical compounds are present.

Chlorophenols (CPs) in wastewater represent a type of pollutant that greatly harms human health and the environment because they are toxic, teratogenic, and carcinogenic [22,23]. Nowadays, CPs are widespread in the environment, even in the most remote natural environments, as well as in aquatic and terrestrial food chains [24]. Among them, 2,4-DCP is present in most wastewaters generated by the textile and the pulp and papermaking industries and has attracted considerable attention because it is highly toxic and difficult to degrade. Furthermore, it is a kind of absorbable organic halide (AOX), which is the main pollution product in the pulping and papermaking industry [25]. It is poisonous, carcinogenic, and teratogenic. Direct discharge into water bodies can cause serious damage to the water environment. As a result, 2,4-DCP is usually degraded by AOPs and used as a target pollutant for developing new AOP methods.

In our previous study, we confirmed that the EDDS-modified photo-Fenton system can effectively degrade 2,4-DCP in deionized water in the laboratory. The effect of pH, H2O2 concentration, and Fe(III)-EDDS dosage was investigated, and the optimal condition was determined [26]. On the other hand, it was found that 2,4-DCP could be effectively degraded in the EDDS-modified photo-Fenton system at pH 3–7. Furthermore, •OH radicals were found to be the main active species of degradation. In this study, our purpose is to reveal the effect of chemical compounds (organic and inorganic compounds) on 2,4- DCP degradation using the EDDS-modified photo-Fenton system and the effect of the complex water matrix on the system. First, the effects of organic matter such as humic acids (HAs) (representing common organic matter in natural water) and 2,4,6-trichlorophenol (2,4,6-TCP) (representing the same kind of organic compounds always accompanying wastewater) on the degradation of 2,4-DCP are discussed. Second, the effects of inorganic ions on 2,4-DCP degradation are evaluated. Finally, three types of natural waters, including natural lake water (NLW), effluent from domestic sewage treatment plants (DSTP), and secondary effluent from pulp and paper mill wastewater (PPMW), are selected as the natural water matrix for 2,4-DCP degradation in the EDDS-modified photo-Fenton reaction. A metal halide lamp is used to mimic sunlight during these laboratory experiments. The expected selectivity of the EDDS-modified photo-Fenton reaction for 2,4-DCP in natural water is demonstrated by comparing the decrease in 2,4-DCP content in pure water with that in different natural waters. The results of these experiments on the efficiency of the EDDS-modified photo-Fenton system in removing pollutants from natural waters have important practical implications for advancing water treatment technologies.

#### **2. Materials and Methods**

#### *2.1. Chemicals*

EDDS (35% in water) was purchased from Shanghai Anpu Experimental Technology Co., Ltd. (Shanghai, China). 2,4-DCP and 2,4,6-TCP were purchased from Shanghai Macklin Biochemical (Shanghai, China). Ferric chloride hexahydrate, sodium chloride, sodium sulfate, sodium carbonate, sodium nitrate, sodium phosphate dodecahydrate, potassium chloride, magnesium chloride, calcium chloride, and H2O2 (30% in water) were purchased from Guangdong Guanghua Sci-Tech Co., Ltd. (Guangzhou, China). Humic acids (HAs) were purchased from Alfa (Shanghai, China). Methyl alcohol (HPLC grade) and acetonitrile (HPLC grade) were purchased from MERCK (Shanghai, China). The pH of the solutions was adjusted with sodium hydroxide (NaOH) and hydrochloric acid (HCl). The ferric carboxylic acid complex solution was prepared by mixing iron and EDDS aqueous solutions in a ratio of 1:1 (and left to stand for more than 1 h to ensure excellent chelation efficiency).

#### *2.2. Analytical Procedures*

Total organic carbon (TOC) was measured using an Analytikjena TOC-VCSN analyzer (Jena, Gremany). An ICS-5000 ion chromatography (IC) unit (Dionex Corporation, Sunnyvale, CA, USA) was used to monitor the release of chloride ions; this unit was equipped with a conductivity detector, an anion self-regenerating suppressor (ASRSTM <sup>300</sup> × 4 mm, Dionex Corporation, Sunnyvale, CA, USA), and the AutoSuppressionTM Recycle Mode. Ultrapure water and 250 mM NaOH were supplied as the eluent to IC at a flow rate of 1 mL min−1. Degradation of 2,4-DCP, 2,4,6-TCP, and a mixture of 2,4-DCP and 2,4,6-TCP was determined using ultra-high-pressure liquid chromatography (Waters ACQUITY UPLC® H-Class, C-18 column, UV detector) (Waters, Milford, MA, USA). The UPLC operating conditions for the target compounds are listed in Table 1.


#### *2.3. Experimental System*

All experiments were performed in a cylindrical Plexiglas container, which was covered with aluminum foil to protect against light and avoid side photochemical processes, placed on a homemade photoreactor (Figure 1). The photoreactor was designed with a cylindrical container, and the lamp with a glass-jacket was fixed through the central axes of the cell. For the experiment, the target pollutant solution and the Fe(III)-EDDS complex solution (0.1 mM) were added to a 1 L beaker. An adequate volume of the mixture was sampled and transferred intoa1L volumetric flask, to which hydrogen peroxide solution (1.0 mM) was added. The 1 L flask containing the reaction solution was placed under agitation in a reactor equipped with a metal halide lamp (continuous spectrum of 290–800 nm), and samples were taken from the reactor at different time intervals. In order to simulate the natural pH, the initial pH value of all the single-effect experiments was set close to 7.0. Since the reaction may have continued after sampling, methanol was added to stop the reaction.

**Figure 1.** The scheme of the photoreactor used in this study.

#### **3. Results**

#### *3.1. Effects of Organic Matter*

3.1.1. Effect of Organic Compounds

Changes in pollutant concentration could significantly influence the degradation efficiency in the photo-Fenton system. Therefore, the influence of different initial concentrations of 2,4-DCP on the photochemical experiment was studied while keeping the other reaction parameters unchanged. The pollutant was almost completely degraded at concentrations of 5 to 20 mg L−1. For 2,4-DCP concentrations equal to or exceeding 20 mg L<sup>−</sup>1, the degradation rate and efficiency decreased as the 2,4-DCP concentration increased (Figure 2). This phenomenon can be interpreted because the number of hydroxyl radicals did not increase proportionally as the pollutant concentration increased [27]. 2,4- DCP exhibited a significant degradation efficiency of around 53% even at the maximum pollutant concentration of 100 mg L−<sup>1</sup> considered in this study. The experimental results show that the EDDS-modified photo-Fenton process had an obvious degradation effect on 2,4-DCP, even at high initial concentrations.

**Figure 2.** 2,4-DCP degradation efficiency for different 2,4-DCP initial concentrations. [Fe(III)-EDDS] = 0.1 mM, [H2O2] = 1.0 mM, and pH = 7.0 ± 0.1.

The toxicity of CPs can be ascribed to the number of Cl atoms on the benzene ring: the more Cl atoms in the chemical formula of CPs, the more toxic organic compounds there are [28]. Moreover, the quantity of Cl atoms on the benzene ring may affect the rate of degradation. Accordingly, we experimentally investigated the degradation of single 2,4,6-TCP and 2,4-DCP, as well as that of a mixture of these two pollutants.

In our photo-Fenton system and the single-substance experiments, the rate of 2,4-DCP pollutant removal was superior by around 65% to those of 2,4,6-TCP (Figure 3).

**Figure 3.** (**a**) EDDS-modified photo-Fenton degradation of 2,4-DCP and 2,4,6-TCP. (**b**) Variation of ln C/C0 with time in 2,4-DCP and 2,4,6-TCP concentration using the EDDS-modified photo-Fenton system. [2,4-DCP] = [2,4,6-TCP] = 20 mg L−1, [Fe(III)-EDDS] = 0.1 mM, [H2O2] = 1.0 mM, pH = 7.0 ± 0.1.

This difference in 2,4-DCP and 2,4,6-TCP degradation rates could be attributed to the number of chlorine atoms on these CPs, possibly because the OH and Cl groups of 2,4-DCP are aligned along the ortho and para directions, and the •OH radicals have the same preference for attack. By contrast, steric hindrance prevents the hydroxylation of 2,4,6-TCP [29]. It was reported that in the heterogeneous photo-Fenton system, the 4-CP removal rate was superior to that of 2,4,6-TCP, indicating that the quantity of Cl atoms significantly influenced the phenolic compounds' degradation rate [30]. However, our results showed the same effect but also that the EDDS-modified photo-Fenton process can be used to effectively treat 2,4-DCP and 2,4,6-TCP pollutants in deionized water.

Given that wastewaters contain multiple organic matters, we performed other experiments with a mixture of two soluble pollutant compounds (2,4-DCP and 2,4,6-TCP). The individual concentrations of 2,4-DCP and 2,4,6-TCP were set to 10 mg L−<sup>1</sup> (total pollutant concentration is 20 mg L−1) while keeping the other experimental conditions unchanged from those employed in the single-substance experiments.

The removal rate of residual pollutants after 20 min of irradiation was considerably slower than that before 20 min of irradiation (Figures 3 and 4). This was primarily ascribed to the low concentration of residual pollutants (around 20% in the single-pollutant experiment and 40% in the mixture of the two pollutants of the initial concentration) after the first 20 min of reaction, resulting in a higher competition reaction of •OH radicals with Fe2+ and H2O2 able to scavenge •OH radicals as well [31–33].

**Figure 4.** *Cont*.

**Figure 4.** EDDS-modified photo-Fenton degradation of a mixture of 2,4-DCP and 2,4,6-TCP in the two-pollutants experiment ((**a**) 2,4-DCP and 2,4,6-TCP degradation efficiency; (**b**) apparent rate constant of degradation of mixed pollutants). [2,4-DCP] = 10 mg L−1, [2,4,6-TCP] = 10 mg L−1, [Fe(III)-EDDS] = 0.1 mM, [H2O2] = 1.0 mM, and pH = 7.0 ± 0.1.

In the two-pollutants experiment, approximately 80% of the two are degraded after 120 min. Compared with the single-pollutant experiment, the removal rate of the two pollutants was lower by around 54% for 2,4-DCP and 24% for 2,4,6-TCP. This result implies that the degradation of pollutants in complex media is slower. Therefore, our results demonstrate that the EDDS-modified photo-Fenton system is very effective for treating a single organic pollutant or a mixture of two organic pollutants.

The apparent rate constant of degradation of mixed pollutants is lower than that of a single pollutant (Figures 3b and 4b), which indicates that 2,4-DCP degradation will be significantly affected in more complex natural media (a detailed analysis of 2,4-DCP degradation in complex natural media is provided in Section 3.3). Compared with a medium containing a single pollutant, a medium containing two organic pollutants will decrease the removal rate and weaken the removal effect. Nevertheless, in this experiment, the main goal (of pollutant removal) was achieved within a reasonable timespan.

#### 3.1.2. Effect of Humic Acids (HAs)

HAs are ubiquitous in aquatic environments [34]. To a large extent, the degradation of organic pollutants is affected by their interaction with dissolved organic matter (DOM, such as HA) in aquatic environments [35]. When DOM absorbs ultraviolet or solar radiation, it forms reactive oxygen intermediates and can also have a shielding effect and inhibits the AOPs [36,37]. DOM can enhance or inhibit the photodegradation rate [38,39]. DOM in water, such as HA, can trap •OH radicals and also produce •OH radicals under irradiation [40]. HA is a potential electron donor and can reduce Fe3+ in the system. For example, HA can significantly improve the degradation effect of pentachlorophenol by promoting the redox cycles of Fe(III) and Fe(II) in the photo-Fenton system [41,42]. Therefore, it is necessary to study the influence of HA on pollutant degradation in natural water. In our study, when HA was present in the solutions, the degradation rate of the EDDS-modified photo-Fenton system decreased marginally (around 10% less after 120 min of irradiation).

Moreover, when the HA concentration was increased from 2 to 5 mg L−1, the 2,4-DCP degradation rate was almost unchanged (Figure 5).

**Figure 5.** Effect of humic acid (HA) on 2,4-DCP degradation. [2,4-DCP] = 20 mg L−1, [Fe(III)-EDDS] = 0.1 mM, [H2O2] = 1.0 mM, and pH = 7.0 ± 0.1.

The lower degradation rate could be ascribed to the fact that HA scavenged •OH radicals. Moreover, HA and the Fe(III)-EDDS complex compete for light absorption because HA is known to absorb sunlight, which reduces the photoredox process of the iron complex [35,43]. The photochemical properties of HA can be ascribed to complex phenomena and are influenced by multiple factors, including its origin and structural characteristics [35,44].

#### *3.2. Effects of Inorganic Ions*

The results above indicate that the EDDS-modified photo-Fenton process is a promising and novel technology that can completely degrade 2,4-DCP. However, industrial wastewaters contain various inorganic ions. These inorganic anions and cations can play complex roles in the EDDS-modified photo-Fenton process.

Cl− and SO4 <sup>2</sup><sup>−</sup> ions could reduce the reaction efficiency by scavenging hydroxyl radicals and competing with the ligand for the complexation of iron ions [45]. The reaction of •OH radicals with SO4 <sup>2</sup><sup>−</sup> ions occurs only at very high concentrations of SO4 <sup>2</sup><sup>−</sup> ions [46]. Indeed, compared with the control experiment, the addition of SO4 <sup>2</sup><sup>−</sup> (at concentrations of 100 to 200 mM) can marginally improve the 2,4-DCP photodegradation rate of the EDDS-modified photo-Fenton system (Figure 6a).

**Figure 6.** Effects of anions and cations on 2,4-DCP degradation. (**a**) Anions; (**b**) cations. [2,4-DCP] = 20 mg L<sup>−</sup>1, [Fe(III)-EDDS] = 0.1 mM, [H2O2] = 1.0 mM, and pH = 7.0 <sup>±</sup> 0.1.

From our results (Figure 6a), we show that the 2,4-DCP degradation efficiency decreased slightly at the chloride concentration up to 7.10 g/L, and the process continued to exhibit a significant degradation efficiency. The addition of chlorine ions to an aqueous solution of iron ions will result in the formation of the Fe(Cl)2+ complex, which has a weaker (photo)reactivity than the Fe(III)-EDDS complex (R1) [47]. The effect of the concentration of chlorine ions is complicated in Fenton chemistry. When the Cl− concentration is equal to 17.75 g/L, the 2,4-DCP degradation efficiency is higher than when a 7.10 g/L Cl− concentration is added and lower than that in the deionized aqueous solution. This

may be because of the increased concentration of chloride ions and the formation of Cl• radicals by the Fe(Cl)2+ complex under irradiation (R2), which contribute toward 2,4-DCP degradation. The activity of Cl• radicals is weaker than that of •OH radicals [48,49]. The inhibitory effect of chloride should also be attributed to the reactivity of •OH radical with Cl− leading to the formation of Cl• radical or ClOH• − and after Cl• radical reacts with Cl− to form the radical Cl2 • − [50,51]. These reactions are significant in the presence of •OH radicals and Cl−.

$$Fe^{3+} + Cl^- \rightarrow Fe(Cl)^{2+} \tag{1}$$

$$\text{Fe(Cl}^{\cdot}\text{)}^{2+} + h\upsilon \rightarrow \text{Fe}^{2+} + \text{Cl}^{\bullet} \tag{2}$$

We found that NO3 − ions had a small effect on the system. When the NO3 − ion concentration was increased, the degradation rate slightly increased (Figure 6a). NO3 − ions can produce other •OH radicals (R3 and R4) under irradiation. However, NO3 − ions have a strong ultraviolet (UV)-shielding effect, which is more significant than the formation of •OH radicals through NO3 − photolysis [52]. The fact that the presence of NO3 − ions did not significantly influence the degradation of 2,4-DCP can possibly be ascribed to the interaction of all the above mentioned factors.

$$NO\_3^- + hv \rightarrow NO\_2^- + O(3P) \tag{3}$$

$$NO\_3^- + H^+ + hv \rightarrow NO\_2^\bullet + HO^\bullet \tag{4}$$

On the contrary, the degradation of 2,4-DCP was severely affected by the presence of carbonates (Figure 6a). When the carbonate concentration was 300 mg/L, the 2,4-DCP degradation efficiency decreased by approximately 25%. It has been reported that carbonates play an essential role in AOPs by acting as scavengers of hydroxyl radicals through the reaction R5, increasing oxidant consumption [53]. Papautsakis et al. [54] reported that carbonate can scavenge •OH radicals and inhibit the degradation of imidacloprid in the Fe-EDDS photo-Fenton process. Moreover, carbonate has also been shown to have a destructive effect on the stability of soluble iron [53].

$$\begin{array}{l}\text{CO}\_3^{2-} + HO^\bullet \rightarrow \text{CO}\_3^{\bullet-} + OH^-\\(k = 3.9 \times 10^8 M^{-1}s^{-1})\end{array} \tag{5}$$

As the phosphate concentration increased, the 2,4-DCP degradation rate decreased, indicating that the PO4 <sup>3</sup><sup>−</sup> ion forms a complex with Fe3+ and reduces the efficiency of the EDDS-modified photo-Fenton system. It is reported in the literature that Fe3+ precipitation by PO4 <sup>3</sup><sup>−</sup> limited the Fenton and photo-Fenton reactions in aqueous solutions with high phosphate concentrations [55].

It is known that pH significantly influences the Fenton process. In the presence of carbonates and phosphates, the solution pH has a buffering capacity, owing to the hydrolysis of PO4 <sup>3</sup><sup>−</sup> and CO3 <sup>2</sup><sup>−</sup> ions. Under this condition, the solution pH is neutral at the end of the experiment, while in deionized water, the corresponding solution pH is approximately 4.8. This may be also an important reason for the effect of carbonate and phosphate ions on the removal of pollutants.

When K+, Ca2+, and Mg2+ ions were present in the solution, the degradation efficiency of the system decreased slightly (Figure 6b). Metal cations affect the degradation of the system by competing with Fe(III) ions for ligands. It has been shown that the higher the number of charges of metal ions, the greater their complexation ability [56]. The complexation ability of Fe3+ ions is the strongest, followed by Ca2+ and Mg2+ ions, and K<sup>+</sup> ions are the weakest. Therefore, even in the case of very high K<sup>+</sup> concentration, the effect on the system is very small.

The above results made clear that the EDDS-modified photo-Fenton process can efficiently degrade organic pollutants in the presence of inorganic ions. Even at high concentrations of inorganic ions, the EDDS-modified photo-Fenton system exhibited significant degradation efficiency. Furthermore, inorganic ions affect degradation mainly by

competing with trivalent iron for ligands and scavenging hydroxyl radicals. Therefore, an investigation of 2,4-DCP degradation using the EDDS-modified photo-Fenton system in the presence of the main constituents of water will help to improve our understanding of their effects on pollutant degradation in real aquatic systems.

#### *3.3. Effect of Natural Water Bodies*

TOC and ionic chromatography analyses of the three natural water bodies show significantly different matrix contents. In terms of ions, the amounts of Cl−, NO3 −, and SO4 <sup>2</sup><sup>−</sup> detected in the DSTP, PPMW, and NLW water samples varied significantly. NO3 − was not detected in NLW (Table 2). The chromaticity of the three water bodies varied considerably, which is of great importance for the photo-Fenton system. In conclusion, the chemical compositions of the three water samples differed considerably, and the influences of ion concentration and chromaticity were non-negligible.

**Table 2.** Physicochemical parameters and chemical compositions of natural water bodies.


Pollutant degradation in natural water is a more complex process than degradation in the presence of single inorganic ions and organic matter. The EDDS-modified photo-Fenton system was used to evaluate the photocatalytic removal of 2,4-DCP dissolved in natural waters, including NLW, DSTP, and PPMW. The 2,4-DCP degradation rate in deionized water was considerably faster than that in natural waters. The degradation efficiencies after 120 min of irradiation were 52.5%, 64.4%, and 38.5% in DSTP, NLW, and PPWM, respectively (Figure 7) and around 100% in deionized water after 60 min of irradiation.

The lower 2,4-DCP degradation rate in NLW, DSTP, or PPMW than that in deionized water could be ascribed to the presence of inorganic ions and dissolved organic compounds. As given in Table 2, we determined some of the chemical constituents. The concentrations of chloride, sulfate, and nitrate ions in the studied natural waters were different. Among the three aforementioned water systems, 2,4-DCP degradation was most severely inhibited in PPMW, which may be ascribed to the relatively higher concentrations of inorganic ions in this compartment than those in the other two other water systems. Secondly, PPMW is dark yellow in color, which means it can cause a screen effect and thus decrease the Fe(III)-EDDS photoredox process. Thirdly, the presence of higher TC concentration can affect the degradation efficiency of 2,4-DCP. A more comprehensive understanding of the influence of TOC and total inorganic ion concentration in water on pollutant removal is therefore needed.

**Figure 7.** Degradation of 2,4-DCP in natural water bodies. [2,4-DCP] = 20 mg L−1, [Fe(III)-EDDS] = 0.1 mM, [H2O2] = 1.0 mM, and pH = 7.0 ± 0.1.

The TOC concentration of NLW, DSTP, and PPMW increased successively, contrary to the decreased 2,4-DCP degradation efficiencies (Figure 8a). The high concentration of TOC in the natural water body was in competition with the target pollutant for the reactivity of •OH radicals, which is an essential reason for the resulting low degradation efficiency. Moreover, compared with that in deionized water, the decreased 2,4-DCP photodegradation efficiency in natural water may be ascribed to the optical filter effect of organic matter in natural water. Indeed, organic matter can be one of the critical absorbers of sunlight in aquatic environments [35]. We speculate that TOC concentration is not the only factor affecting 2,4-DCP degradation in the studied water bodies. The effect of total inorganic ion concentration in water on 2,4-DCP removal was explored. Overall, the 2,4-DCP degradation efficiency decreased as the total inorganic ion concentration increased (Figure 8b). The same effect was observed in the study by Sakkas et al. [25]. As the salinity of water increased, chlorothalonil combined with DOM through hydrophobic interaction or weak van der Waals forces, thus affecting the photodegradation of the pollutants. Obviously, the factors affecting the 2,4-DCP removal in natural water included inorganic ion concentration, TOC concentration, and chromaticity.

The UV–visible absorption spectrum of different water matrices containing 2,4-DCP is shown in Figure 9, and it was noted that most change in UV–visible absorption occurred in the UV zone (lower than 290 nm), which was out of the wavelength range of the lamp used in this study. As a result, it was preliminarily indicated that the influence had nothing to do with UV–visible absorption.

**Figure 8.** Effects of (**a**) TOC and (**b**) total inorganic ion concentration on 2,4-DCP degradation efficiency in different water bodies. [2,4-DCP] = 20 mg L<sup>−</sup>1, [Fe(III)-EDDS] = 0.1 mM, [H2O2] = 1.0 mM, and pH = 7.0 ± 0.1.

**Figure 9.** UV-visible absorption spectrum of different water bodies.

The degradation of 2,4-DCP was affected differently by different natural water compositions. However, the more the water is loaded with organic/inorganic compounds, the more significant the inhibition of 2,4-DCP degradation. Regardless, the EDDS-modified photo-Fenton system still removed more than 50% of 2,4-DCP in NLW and DSTP and more than 30% in PPMW. This finding indicates that the EDDS-modified photo-Fenton process can be used to effectively treat pollutants in natural waters and potentially simulate solar photocatalytic water treatment. Thus, 2,4-DCP degradation in natural water bodies justifies a more in-depth study to understand and evaluate the parameters that are essential for the efficiency of the process.

#### **4. Conclusions**

The results indicate that the EDDS-modified photo-Fenton system is suitable for 2,4- DCP removal. It is a promising route for treating 2,4-DCP by simulating natural sunlight, which is a low-cost alternative light source and significantly reduces the process cost. The system could effectively degrade single 2,4-DCP and 2,4,6-TCP pollutants and the mixture of 2,4-DCP and 2,4,6-TCP. Furthermore, it could effectively degrade pollutants in the presence of common inorganic ions. The effect of anions on 2,4-DCP degradation was found to be stronger than that of cations. Finally, the degradations of 2,4-DCP in different water bodies, including NLW, DSTP, and PPMW, were remarkably different. The 2,4-DCP degradation rate in PPMW was severely inhibited, which may be related to the high absorption of light, high TOC concentration, and high inorganic ion content in this water compartment. The use of several different natural waters to treat 2,4-DCP with the EDDS-modified photo-Fenton process shows the efficiency of this process for industrial applications. Nevertheless, the application will be particularly more efficient when this process is used in a ternary treatment when the water is not too loaded.

**Author Contributions:** Conceptualization, W.H. and G.M.; methodology, W.H. and Y.H.; software, Y.H.; validation, W.H. and S.W.; formal analysis, W.H. and G.M.; investigation, W.H. and Y.H.; resources, G.M.; data curation, G.M.; writing—original draft preparation, W.H.; writing—review and editing, G.M.; visualization, H.L.; supervision, G.M.; project administration, W.H. All authors have read and agreed to the published version of the manuscript.

**Funding:** This research received no external funding.

**Institutional Review Board Statement:** Not applicable.

**Informed Consent Statement:** Not applicable.

**Acknowledgments:** This work was partially supported by the National Natural Science Foundation of China (No. 21367003), Guangxi Science and Technology Research Program (No. AA17202032) and Open Fond of Guangxi Key Laboratory of Clean Pulp & Papermaking and Pollution Control (KF201724).

**Conflicts of Interest:** The authors declare no conflict of interest.

#### **References**


## *Article* **Computational Fluid Dynamics Modeling of Rotating Annular VUV/UV Photoreactor for Water Treatment**

**Minghan Luo 1,2,\*, Wenjie Xu 1, Xiaorong Kang 1, Keqiang Ding <sup>1</sup> and Taeseop Jeong <sup>3</sup>**


**\*** Correspondence: leon96201@njit.edu.cn; Tel.: +86-25-86118963

**Abstract:** The ultraviolet photochemical degradation process is widely recognized as a low-cost, environmentally friendly, and sustainable technology for water treatment. This study integrated computational fluid dynamics (CFD) and a photoreactive kinetic model to investigate the effects of flow characteristics on the contaminant degradation performance of a rotating annular photoreactor with a vacuum-UV (VUV)/UV process performed in continuous flow mode. The results demonstrated that the introduced fluid remained in intensive rotational movement inside the reactor for a wide range of inflow rates, and the rotational movement was enhanced with increasing influent speed within the studied velocity range. The CFD modeling results were consistent with the experimental abatement of methylene blue (MB), although the model slightly overestimated MB degradation because it did not fully account for the consumption of OH radicals from byproducts generated in the MB decomposition processes. The OH radical generation and contaminant degradation efficiency of the VUV/UV process showed strong correlation with the mixing level in a photoreactor, which confirmed the promising potential of the developed rotating annular VUV reactor in water treatment.

**Keywords:** VUV; photoreactor; CFD; MB; water treatment

#### **1. Introduction**

Use of ultraviolet-based photoreactors in water-treatment processes is rapidly increasing, and ultraviolet-based advanced oxidation processes (UV AOPs) have been studied for over 30 years. The H2O2/UV process presents increased economic cost and technical complexity due to the treatment of residual peroxide, leading to its application only in small and medium-sized water treatment facilities. The VUV/UV process uses ozone-generating mercury lamps that emit 185 nm VUV and 254 nm UV radiation, in which the 185 nm radiation reacts with water to produce hydroxyl radicals (·OH). Therefore, VUV/UV photodegradation is considered to be a simple and environmentally friendly water-treatment technology with attractive economic potential, which has shown promising potential in wastewater treatment [1]. Although plenty of experiments have yielded promising results at lab-scale, the VUV/UV AOP has not yet been implemented at a full-scale plant in water treatment. There are still problems that impede large-scale application of VUV/UV photoreactors in the water-remediation field. For example, lack of a proper simulation model to predict and analyze the performance of VUV/UV photoreactors is one of the problems hindering their practical implementation. An effective modeling of the VUV/UV process involves the simultaneous solution of momentum equations, mass transfer equations, and radiation energy equations (UV and VUV radiations), along with a complex kinetic scheme of more than 40 reactions.

Computational fluid dynamics (CFD) is an established and effective tool for modeling complex fluid dynamic processes, and has been used extensively for the design, optimization, and scale-up of UV disinfection and oxidation photoreactors in recent years [2,3].

**Citation:** Luo, M.; Xu, W.; Kang, X.; Ding, K.; Jeong, T. Computational Fluid Dynamics Modeling of Rotating Annular VUV/UV Photoreactor for Water Treatment. *Processes* **2021**, *9*, 79. https://doi.org/10.3390/pr9010079

Received: 1 December 2020 Accepted: 29 December 2020 Published: 31 December 2020

**Publisher's Note:** MDPI stays neutral with regard to jurisdictional claims in published maps and institutional affiliations.

**Copyright:** © 2020 by the authors. Licensee MDPI, Basel, Switzerland. This article is an open access article distributed under the terms and conditions of the Creative Commons Attribution (CC BY) license (https:// creativecommons.org/licenses/by/ 4.0/).

Previous studies have pointed out the importance of using a comprehensive kinetic scheme and a detailed radiation model, including the reflection, refraction, and absorption of photons in UV photoreactors, for CFD simulation. However, few studies about modeling of H2O2/UV using the CFD method have individually reported the role of 254 nm irradiation in direct photolysis of water, or of the ·OH radical oxidation pathways in the process of removing target pollutants [2,4,5]. Moreover, despite of the similarities between the H2O2/UV and VUV/UV processes, the associated hydroxyl radical generation mechanisms are different [4–6]. The production of ·OH radicals in VUV/UV systems relies on the photolysis of water at 185 nm irradiation, while hydrogen peroxide photolysis at 254 nm irradiation is the predominant mechanism for the generation of ·OH radicals in H2O2/UV systems [5]. The direct photolysis of water will generate species such as ·OH, ·H and H+, whereas the UV photolysis of hydrogen peroxide produces only ·OH, therefore causing different radical reaction schemes during H2O2/UV and VUV/UV processes. In addition, the emissions at 185 nm and 254 nm synchronously contribute to the removal of contaminants in a VUV/UV process. In contrast, H2O2/UV approaches primarily rely on the degradation functions of 254 nm photons. UV light at a wavelength of 185 nm, which plays the key role in ·OH production in VUV/UV AOPs, is transmitted a relatively short distance in solutions. As a result, VUV/UV AOPs normally require better mixing within reactors than UV AOPs, and the effective identification of mixing characteristics of the area around UV lamps is therefore of particular significance for VUV/UV AOP studies.

In this context, this work aimed to develop a comprehensive CFD simulation tool able to make an in-depth analysis of the VUV/UV process applied to water treatment. The proposed computational model integrates a series of sub-models such as hydrodynamic simulations, a multispecies mass transport model, chemical reaction kinetics, and irradiance distribution within the reactor. The radiation field within the reactor was modeled using a nongray discrete ordinate (DO) sub model, which allowed for independent and simultaneous studies of the transportation paths of 185 nm VUV and 254 nm UV. The developed model was experimentally evaluated in a continuous-flow VUV/UV photoreactor for the treatment of a selected pollutant: methylene blue (MB). Finally, we establish and discuss a model for degradation pathways within VUV/UV photoreactors. The results from this study revealed crucial hydrodynamic characteristics in the VUV/UV photoreactor, and provide useful suggestions for the design and optimization of VUV/UV photoreactors, promoting the practical application of VUV/UV techniques in the water-treatment field.

#### **2. Materials and Methods**

#### *2.1. Hydrodynamics*

Based on the principles of conservation of mass and momentum, the continuity equations in a rotating annular VUV reactor (RAVR) were described. A three-dimensional computational fluid dynamic model was developed to calculate the local hydrodynamics in the photoreactor.

$$\sum\_{i=1}^{n} a\_i = 1 \tag{1}$$

where *n* is the total number of phases; the subscript *i* represents the gas or liquid phase. The conservation equations are written by performing an ensemble average of the local instantaneous balance for each phase. The motion of each phase is governed by the corresponding mass and momentum conservation equations.

Continuity equation:

$$\frac{\partial(\boldsymbol{a}\_{i}\cdot\boldsymbol{\rho}\_{i})}{\partial\mathbf{t}} + \nabla\cdot(\boldsymbol{a}\_{i}\boldsymbol{\rho}\_{i}\stackrel{\rightarrow}{\boldsymbol{u}}\_{i}) = \boldsymbol{0} \tag{2}$$

where *α*, *ρ*, and <sup>→</sup> *u* stand for the volume fraction, density, and velocity vector, respectively. Momentum equation:

$$\frac{\partial(a\_i\rho\_i\stackrel{\rightarrow}{\boldsymbol{u}\_i})}{\partial\mathbf{t}} + \nabla\cdot(a\_i\rho\_i\stackrel{\rightarrow}{\boldsymbol{u}\_i}\stackrel{\rightarrow}{\boldsymbol{u}\_i}) = -a\_i\nabla P\_i + \nabla\cdot\left(a\_i\mu\_i(\boldsymbol{\nabla}\stackrel{\rightarrow}{\boldsymbol{u}\_i} - (\boldsymbol{\nabla}\stackrel{\rightarrow}{\boldsymbol{u}\_i})^T)\right) + a\_i\rho\_i\stackrel{\rightarrow}{\boldsymbol{g}} \pm \stackrel{\rightarrow}{F}\_i \tag{3}$$

where *P*, *μ*, and <sup>→</sup> *<sup>g</sup>* are the pressure, viscosity, and gravity acceleration, respectively. <sup>→</sup> *Fi* is the interfacial force acting on phase *i* due to the presence of the other phase, *j*.

The turbulent dispersion force is the result of the turbulent fluctuations of liquid velocity. In this study, the standard *k* − *ε* model for single-phase flows was extended for the two-phase flows to simulate the turbulence, which can be described as follows:

$$\frac{\partial}{\partial t}(\mathbf{a}\_{l}\rho\_{l}\mathbf{k}\_{l}) + \frac{\partial}{\partial \mathbf{x}\_{l}}(\mathbf{a}\_{l}\rho\_{l}\stackrel{\rightarrow}{\mathbf{u}}\_{l}\mathbf{k}\_{l}) \quad = \ \frac{\partial}{\partial \mathbf{x}\_{l}}\Big[\mathbf{a}\_{l}\Big(\mu\_{l} + \frac{\mu\_{ll}}{\sigma\_{k}}\Big)\frac{\partial}{\partial \mathbf{x}\_{i}}\mathbf{k}\_{l}\Big] + \mathbf{a}\_{l}\rho\_{l} - \mathbf{a}\_{l}\rho\_{l}\varepsilon\_{l}\tag{4}$$

$$\frac{\partial}{\partial t}(\mathbf{a}\_{l}\boldsymbol{\rho}\_{l}\boldsymbol{\varepsilon}\_{l}) + \frac{\partial}{\partial \mathbf{x}\_{l}}(\mathbf{a}\_{l}\boldsymbol{\rho}\_{l}\overset{\rightarrow}{\boldsymbol{u}}\_{l}\boldsymbol{\varepsilon}\_{l}) \quad = \ \frac{\partial}{\partial \mathbf{x}\_{l}}\Big[\mathbf{a}\_{l}\Big(\boldsymbol{\mu}\_{l} + \frac{\boldsymbol{\mu}\_{ll}}{\boldsymbol{\sigma}\_{l}}\Big)\frac{\partial}{\partial \mathbf{x}\_{l}}\boldsymbol{\varepsilon}\_{l}\Big] \\ + \boldsymbol{a}\_{l}\frac{\boldsymbol{\varepsilon}\_{l}}{k\_{l}}(\mathbf{C}\_{\boldsymbol{\varepsilon}1p\_{l}} - \mathbf{C}\_{\boldsymbol{\varepsilon}2p\_{l}\boldsymbol{\varepsilon}\_{l}}) \quad \text{(5)}$$

where *Cε*1, *Cε*2, *σk*, and *σε* are parameters in the standard k − ε model and the following values were selected: *Cε*<sup>1</sup> = 1.44, *Cε*<sup>2</sup> = 1.92, *σ<sup>k</sup>* = 1.0, and *σε* = 1.3. In addition, the turbulent viscosities *μtl* can be computed by other equations [7,8].

#### *2.2. Radiative Transfer Model*

The radiative transfer equation (RTE) for an absorbing, emitting, and scattering medium at position <sup>→</sup> *r* in the direction <sup>→</sup> *s* is as follows:

$$\frac{dI(\stackrel{\rightarrow}{r},\stackrel{\rightarrow}{s})}{ds} + (a+\sigma\_{\mathrm{s}})\ I(\stackrel{\rightarrow}{r},\stackrel{\rightarrow}{s}) = an^{2}\frac{\sigma T^{4}}{\pi} + \frac{\sigma\_{\mathrm{s}}}{4\pi} \int \stackrel{4\pi}{l} I(\stackrel{\rightarrow}{r},\stackrel{\rightarrow}{s}') \otimes \left(\stackrel{\rightarrow}{s},\stackrel{\rightarrow}{s}'\right) d\Omega' \tag{6}$$

where <sup>→</sup> *r* and <sup>→</sup> *s* are position and direction vectors, respectively. *I* is the radiation intensity, which depends on position and direction; *n* is the refractive index; σ is the Stefan– Boltzmann constant (5.67 × <sup>10</sup>−<sup>8</sup> Wm−2K−4); <sup>α</sup> is the absorption coefficient; *<sup>σ</sup><sup>s</sup>* is the scattering coefficient; ∅ is the phase function; and Ω is the solid angle. Additionally, (*a* + *σs*)s is the optical thickness or opacity of the fluid (water mixture). The refractive index *n* is important when considering radiation in semitransparent media [7].

#### *2.3. Kinetic Reaction Model*

In previous studies, detailed kinetic models for VUV systems have been studied to find the perfect batch-scale mixing conditions [4,5]. In this study, 26 types of reaction occurring in the VUV/UV photoreactor (i.e., equilibrium, photochemical, and radical reactions) were summarized, as shown in Table 1, with reference to the previous study. Thus, in the presence of VUV and UV radiation, the main degradation pathways of the species are initiated by the OH radicals produced by the decomposition of water by 185 nm radiation, leading to radical chain reactions induced by 185 and 254 nm radiation.


**Table 1.** Kinetic model of the VUV/UV photoreactor for degradation of MB.

#### *2.4. Geometry of RAVR and System Setup*

A simple three-dimensional geometry and a mesh structure developed for the RAVR are shown in Figure 1a. The geometry was created using ANSYS DesignModeler software. The RAVR consisted of a reactor with a total length of 500 mm, a 20 mm diameter lamp and a 20 mm diameter inlet, a 30 mm diameter exit tube, and a 5 mm wall thickness. The inlet exit tube was attached to the reactor in a tangential direction to increase the mixing and reactivity of the reactor. The inlet and outlet entered the reactor tangentially to induce rotational flow in the reactor. The RAVR volume was discretized into 132,529 structured and unstructured volume cells using ANSYS Meshing software.

A flow-through, continuously operating RAVR was used to experimentally evaluate the CFD results. The RAVR system is presented in Figure 1b. The reactor had a tangential inlet and an outlet, with annular and rotation flow configuration, operated with a 17 W low-pressure mercury lamp (G10T5VH, Light Sources Inc., Orange, CT, USA) longitudinally placed at the axial center of the reactor. The inlet flow rate was adjusted to within a 1.963–23.550 L/min range with a defined concentration of MB, and the hydraulic retention time varied from 10 s to 120 s. Flow rates of the MB solutions were controlled by peristaltic pumps.

**Figure 1.** The rotating reactor geometry (**a**) and the VUV reactor system scheme (**b**).

#### *2.5. Boundary Conditions and Numerical Solution*

The inlet velocity had a range of 0.104–1.250 m/s, which corresponded to a flow rate of 1.963 to 23.550 L/min. The inlet concentration of the model contaminant MB was equal to 0.5 ppm. A no-slip boundary condition was imposed on the walls. In addition, zero diffusive flux of species was specified at the walls. As per radiation field boundary conditions, the radiation of lamp was defined as a zero-thickness, semitransparent, nonreflecting wall. The density and viscosity of water considered were 998.2 kg/m<sup>3</sup> and 1.003 × <sup>10</sup>−<sup>3</sup> Pa, respectively. The refractive indexes of 185 nm and 254 nm were assigned as 1.458 and 1.376, respectively, and the absorption coefficients 35.67 (m<sup>−</sup>1, UVT = 70%) and 12.78 (m−1, UVT = 88%), respectively.

ANSYS 16.2 Fluent was employed to read the mesh and perform the CFD computations. The segregated steady-state solver was used to solve the governing equations. Second-order upwind discretization schemes were applied except for pressure, for which the standard scheme was selected. The semi-implicit method for pressure linked equations (SIMPLE) algorithm was chosen for the pressure–velocity coupling. The variation of velocity magnitude, model contaminant concentration, and irradiation flux at several points of the computational domain were used as indicators of convergence (at least 20 iterations). Additionally, convergence of the numerical solution was assured by monitoring the scaled residuals to a criterion of at least 10−<sup>4</sup> for the concentration of MB. While the simulation was always tracked with time, the solution algorithm was run with both steady and transient flow simulations.

#### *2.6. Chemicals and Analytical Methods*

For the VUV reactor experiments, the chemicals used for experiments were reagentgrade or higher, supplied by Sigma-Aldrich. MB powder was used as purchased without further purification. The MB was diluted separately with ultrapure laboratory water. Distilled water was used in all experiments and analytical determinations. The concentration of MB in the VUV reactor effluent was determined spectrophotometrically following the peak at 664 nm using a UV spectrophotometric probe (UV1800, Shimadzu Co., Marlborough, MA, USA, Kyoto, Japan spectrophotometer). The concentration of hydrogen peroxide was determined via UV spectrophotometry utilizing the *I* − <sup>3</sup> method [16].

#### **3. Results and Discussion**

#### *3.1. Hydrodynamics*

The fluid velocity magnitudes are shown in Figure 2 for selected cross-sections in the designed RAVR. The results showed that the fluid entered the inlet at a high velocity and initiated a rotational motion along the wall of the RAVR. For the radial velocity distribution of the reactor, the velocity on the wall of the reactor was higher than that on the surface of central light source of the reactor. Because the rotational flow flowed tangentially onto the reactor wall, part of the kinetic energy was consumed when it reached the light source.

**Figure 2.** Velocity vectors (**a**) and contours of velocity magnitude (**b**) along the axial plane at different section planes located at 2 (P-1), 12.5 (P-2), 25 (P-3), 36.5 (P-4), and 48 (P-5) cm positions from the inlet.

As shown by the streamline in Figure 3, the fluid maintained rotational motion from the inlet to the outlet of the reactor, and thus the outflow also presented a rotating flow. Obviously, the rotational flow in the vicinity of outer wall of the reactor showed an upward motion; the rotation force was generated from the flowing force of fluid entering the inlet without additional energy supply. The fluid from the inlet raised along the reactor wall with a high movement velocity. The RAVR integrated the flow characteristics of a continuous-flow stirred-tank reactor (CSTR) and a plug flow reactor (PFR). As mentioned in Section 1, the RAVR is an economical reactor with good mixing function compared with long tubular reactors [17–19] or reactors with internal baffles [3,20,21]. The rotating flow along the VUV lamp in a RAVR reactor plays the role of mixing and extending the fluid-retention time, and there is thus no dead zone in a RAVR reactor.

#### *3.2. Pressure Field*

Pressure contour analyses of the longitudinal section of the reactor and of Plane-1 to Plane-5 were conducted for the 3.925 L/min inflow rate, and the results are plotted in Figure 4. Pressure increased from the inlet area to outlet area, and decreased along the radial direction from the reactor wall to the central lamp. For Plane-1 (P-1), where the rotation force of the fluid was induced, the entrance position and out-wall area presented high pressure, indicating a whirl flow along the wall surface. The pressure distribution on the cross-section (*X*–*Y* plane) of the reactor showed a low-pressure zone in the forced vortex region at the central position due to a high fluid rotating velocity. Since fluid flows from a high-pressure area to a low-pressure area, the longitudinal and radial distribution of pressure in reactor revealed that two circulation flows formed in the reactor: (i) an upward rotating flow along the out-wall surface and (ii) a downward stream adhering to the UV lamp from the outlet position to the inlet position. Moreover, the pressure distribution shown in Figure 4 is consistent with the streamline shown in Figure 3. When the fluid entered the reactor, it climbed up along the reactor out-wall surface in a rotating flow (outer circulation ring) and flowed into the central part of reactor due to the pressure difference, then went down along the UV lamp at a slow flow rate (inner circulation ring). In this way, the retention time of introduced solutions in the RAVR reactor was extended through two circulation flows, thus enhancing the treatment effect. Additionally, the pressure difference of the RAVR system, which was identified by the ratio of maximum pressure to minimum pressure, was found to change with inflow rates.

**Figure 3.** Streamline along the axial plane for inflows at 3.952 L/min velocity.

**Figure 4.** The contour and profiles of pressure in the axial direction of the reactor (including Plane-1 to Plane-5) at 3.925 L/min inflow rate.

#### *3.3. UV Radiation*

Figure 5 shows the contour distribution of ultraviolet radiation intensity on the *X–Z* and *X–Y* cross-sections of the reactor. The ultraviolet radiation at 185 nm, which reacts with and causes the formation of ·OH radicals, was decreased to almost zero in the wall area around 4 cm away from the lamp due to the limited transmission ability of ultraviolet radiation at 185 nm in water.

**Figure 5.** Local values of ultraviolet irradiance calculated in the whole reaction zone in the *X–Z* (**c**) and *X–Y* (**d**) planes of the reactor. The longitudinal (**a-1**, **b-1**) and radial (**a-2**, **b-2**) contours of lamp irradiance. a, 185 nm; b, 254 nm.

#### *3.4. Degradation Reaction*

The prediction of the RAVR performance was based on CFD modeling simulations, including the specific chemical kinetics of the reactions in the mass balance of involved species. On the basis of the hydrodynamic distribution characteristics of the RAVR, the kinetic responses of the VUV–H2O–MB reactions in Table 1 at 3.925 L/min inflow rate are represented in Figure 6, with chemical reaction rates calculated by user defined function (UDF).

**Figure 6.** Concentration profile (mol/L) of the species in the center line of the plane as a function of the radial position in the reactor (flow rate = 3.925 L/min).

The profiles of molar concentrations of ·OH radicals (Figure 6) showed a gradual increase from the inlet position to the outlet position (Plane-1 to Plane-4), and then a decrease to a low level on Plane-5. As the Plane-1 was located at the inlet with a small dose of ultraviolet radiation, the concentration of ·OH radicals were accordingly low. More ·OH radicals were generated from P-2 to P-4 with increased ultraviolet radiation. Similarly, the amount of radicals was decreased from the center to wall area in the reactor due to the diminished UV radiation. Moreover, comparing the change of ·OH radicals on different cross-sections, the radical decrease rate from the UV lamp (center position) to the wall surface was slowed down for Plane-2 to Plane-4. In other words, the concentration gradient of ·OH radicals on the same cross-section was decreased along the axial direction from inlet to outlet in the RAVR, albeit not as much as in P-1 and P-5. This is mostly because the upward flow accumulated ·OH radicals and weakened the concentration gradient. It should be noted here that the presented concentration changes of ·OH radicals in Figure 6 do not reflect the quantitative amounts of ·OH radicals directly generated from water by UV radiation: in fact, the radiation at 185 nm has a very low ultraviolet transmittance (UVT) and reaches almost zero at a distance of about 3–4 cm from the lamp source. Thus, the concentration profile of hydroxyl radicals reported in Figure 6 is rather a quantitative result based on the advection and diffusion of fluid in the reactor. In addition, the same distribution feature of ·OH radicals was observed for species like H2O2, ·H, ·HO2, and ·O<sup>−</sup> 2 , except that the concentrations of ·H on Plane-2 to Plane-4 were very similar. As shown in Table 1, the reaction characteristics of the H radical are quite different from those of other species with higher kinetic parameters. Therefore, a gradient concentration along the radial direction was observed for the H radical at the central position of the reactor (P-2, P-3, and P-4), with higher concentrations near the light source. Furthermore, taking into account that the P-2, P-3, and P-4 profiles fully overlapped, it is possible to state that no appreciable gradient concentration occurred along the longitudinal direction in the central part of the reactor. On the other hand, the MB concentration showed an opposite trend to that of ·OH radicals as MB was decomposed and removed by the ·OH radicals. The ·OH-radical-rich zone coincided with active decomposition reactions of MB. Generally, the low concentration of MB in outflow confirmed the good mixing properties of the designed RAVR, which also implies that developing a process with excellent blending function to enhance the contact between ultraviolet light and each reactant is one important route to increase the efficiency of photolysis reactions.

#### *3.5. Comparison of Experimental and Simulation Results*

There are two major concerns to be faced in improving the photochemical reactions of VUV/UV AOP. First, the water must be sufficiently irradiated with 185 nm ultraviolet light for the production of ·OH radicals. In this study, the ·OH radicals were efficiently generated near the UV lamp (which has a high intensity of ultraviolet radiation at 185 nm) and then quickly diffused across the whole reaction tank. Second, the efficiency of VUV/UV AOP can be significantly enhanced by increasing contact opportunities between the species (like MB) to be treated and the ·OH radicals generated in the chain reactions. Various methods have been developed based on the two mentioned routes. For example, a pipe-type reactor with a small diameter has been explored for maximization of ultraviolet light intensity [17–19,22]. However, this method has the disadvantage that the long contact time (residence time) required for the reaction can be only obtained by designing a long reaction pipe, which is normally limited by actual situations. Another commonly studied method is the installation of baffles to increase vortexes inside the reactor [5,21]. However, this is suitable only for small-scale reactors (like the lab scale) and requires further considerations for application at the industry scale. The adsorption method using an adsorption medium (catalyst) has also been widely studied [21]. Its problem is that the catalysts utilized may hinder the irradiation transmission of ultraviolet light. In conclusion, the design optimization of photoreactors plays a key role in promoting the application of photochemical processes based on ·OH radical generation in water-treatment or wastewater-treatment fields. The study followed this idea and investigated the possibility of maximizing the utilization of rotational force of the fluid in the designed photoreactor by adjusting the inflow velocity of treated solutions.

The predicted and experimental degradation efficiencies of MB and CFD showed good consistency for the 3.925 L/min inflow rate, as shown in Figure 7. In addition, the pollutant abatement calculated by CFD modeling was slightly lower than the experimental value for all tested inflow velocities.

**Figure 7.** Comparison of removal rates of MB of experimental data and CFD-predicted results.

#### **4. Conclusions**

This study aimed to develop a novel rotating annular photoreactor with a tangential inlet and outlet to improve the performance of a VUV AOP for increased degradation efficiency of the photoreactor. The flow characteristics as well as the fluid dynamics were investigated and the kinetic model of involved species was simulated to evaluate the reaction characteristics in the designed reactor. Meanwhile, the concentration profiles of ·OH radicals, target pollutant (MB), and other important reacting species were also determined to assess the reactor properties. The results showed that the introduced fluid was in strong rotational movement inside the reactor across the wide range of influent velocities in this study. Moreover, the rotational movement was enhanced with the increasing of inflow rates in the studied velocity range. The CFD modeling results corresponded well to the experimental degradation of MB. They slightly underestimated degradation due to the limited kinetic analysis about radical annihilation effect from by-products of MB degradation. The results from this study confirmed that the ·OH radical generation and contaminant degradation efficiency of a VUV/UV process showed strong correlation with the mixing degree in a photoreactor. Therefore, the developed RAVR has high potential to promote the scale-up of VUV/UV AOP systems.

**Author Contributions:** Conceptualization, M.L. and T.J. Data curation, W.X. and X.K. Formal analysis, M.L. Funding acquisition, M.L. Methodology, M.L., W.X., X.K., T.J. and K.D. Validation, M.L. Visualization, M.L. Writing-original draft, M.L. Writing-review & editing, M.L, K.D. and T.J. Project administration, M.L. and T.J. Software, X.K. and W.X Experiments. All authors have read and agreed to the published version of the manuscript.

**Funding:** This work was funded by Natural Science Research of Jiangsu Higher Education Institutions of China (No.18KJB610006), and Introduction Talent Scientific Research Foundation Project of NanJing Institute of Technology (No. YKJ201847) and Supported by the Cooperation Fund of Energy Research Institute, Nanjing Institute of Technology (No. CXY201925).

**Conflicts of Interest:** The authors declare no conflict of interest.

#### **References**


MDPI

St. Alban-Anlage 66 4052 Basel Switzerland Tel. +41 61 683 77 34 Fax +41 61 302 89 18 www.mdpi.com

*Processes* Editorial Office E-mail: processes@mdpi.com www.mdpi.com/journal/processes

MDPI St. Alban-Anlage 66 4052 Basel Switzerland

Tel: +41 61 683 77 34 Fax: +41 61 302 89 18

www.mdpi.com

ISBN 978-3-0365-3436-7