*Article* **Doping of TiO<sup>2</sup> Using Metal Waste (Door Key) to Improve Its Photocatalytic Efficiency in the Mineralization of an Emerging Contaminant in an Aqueous Environment**

**Dany Edgar Juárez-Cortazar <sup>1</sup> , José Gilberto Torres-Torres <sup>1</sup> , Aracely Hernandez-Ramirez <sup>2</sup> , Juan Carlos Arévalo-Pérez <sup>1</sup> , Adrián Cervantes-Uribe <sup>1</sup> , Srinivas Godavarthi <sup>1</sup> , Alejandra Elvira Espinosa de los Monteros <sup>1</sup> , Adib Abiu Silahua-Pavón <sup>1</sup> and Adrián Cordero-Garcia 1,\***

	- <sup>2</sup> Facultad de Ciencias Químicas, UANL, Universidad Autónoma de Nuevo León, Ciudad Universitaria, San Nicolás de los Garza 66451, Nuevo León, Mexico; aracely.hernandezrm@uanl.edu.mx
	- **\*** Correspondence: adrian.cordero@ujat.mx; Tel.: +52-993-591-0296

**Abstract:** Photocatalysis is an effective advanced oxidation process to mineralize recalcitrant contaminants in aqueous media. TiO<sup>2</sup> is the most used photocatalyst in this type of process. To improve the deficiencies of this material, one of the most used strategies has been to dope TiO<sup>2</sup> with metallic ions. Chemical reagents are often used as dopant precursors. However, due to the depletion of natural resources, in this work it was proposed to substitute chemical reagents and instead use a metallic residue (door key) as a doping precursor. The materials were synthesized using the sol–gel method and calcined at 400 ◦C to obtain the crystal structure of anatase. The characterization of the materials was carried out using X-ray diffraction (XRD), transmission electron microscopy (TEM), diffuse reflectance spectroscopy (DRS), scanning electron microscopy–energy-dispersive X-ray analysis (SEM-EDX) methods X-ray photoelectron spectroscopy (XPS), and inductively coupled plasma optical emission spectroscopy (ICP-OES). The results obtained indicate that Cu+/Cu2+ and Zn2+ ions coexist in the support, which modifies the physicochemical properties of TiO<sup>2</sup> and improves its photocatalytic efficiency. The synergistic effect of the dopants in TiO<sup>2</sup> allowed the mineralization of diclofenac in an aqueous medium when T-DK (1.0) was used as photocatalyst and simulated solar radiation as an activation source.

**Keywords:** emerging contaminant; photocatalysis; TiO<sup>2</sup> doped; discarded metal waste; sol–gel

### **1. Introduction**

Heterogeneous photocatalysis is an advanced oxidation process that has proven to be effective for the mineralization of recalcitrant contaminants in an aqueous medium. The most efficient photocatalyst for the photocatalytic process is TiO<sup>2</sup> since it has optical, structural, and organoleptic properties, making it efficient in the mineralization of contaminants in an aqueous medium. However, this semiconductor has the disadvantage of having an Eg 3.2 eV, which means that it cannot be activated with UV light, so it absorbs only 5% of solar radiation [1–3]. This restricts the use of solar radiation as an activation source.

Various strategies have been developed to improve the spectral response in the visible region of photocatalytic oxides. One of the most widely used strategies has been to alter the physicochemical properties of oxide by incorporating metallic ions into its crystalline structure [4,5]. It has been shown that metal doping of TiO<sup>2</sup> causes the formation of

**Citation:** Juárez-Cortazar, D.E.; Torres-Torres, J.G.; Hernandez-Ramirez, A.; Arévalo-Pérez, J.C.; Cervantes-Uribe, A.; Godavarthi, S.; de los Monteros, A.E.E.; Silahua-Pavón, A.A.; Cordero-Garcia, A. Doping of TiO<sup>2</sup> Using Metal Waste (Door Key) to Improve Its Photocatalytic Efficiency in the Mineralization of an Emerging Contaminant in an Aqueous Environment. *Water* **2022**, *14*, 1389. https://doi.org/10.3390/ w14091389

Academic Editors: Huijiao Wang, Dionysios (Dion) Demetriou Dionysiou and Yujue Wang

Received: 30 March 2022 Accepted: 21 April 2022 Published: 26 April 2022

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**Copyright:** © 2022 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/).

new energy levels below the conduction band and delays the rapid recombination of electron–hole pairs, which increases the photocatalytic activity of this semiconductor [6,7].

To synthesize TiO<sup>2</sup> doped materials, transition metal salts are used. However, based on analysis by the World Metal Reserve, production from a mine has a shorter timescale than mineral deposit formation, suggesting that known primary metal supplies will be depleted. Importantly, the risk of contamination by heavy metals is higher since they are not chemically or biologically degradable. Once disposed of, they can remain in the environment for hundreds of years and cause environmental damage. It has been shown that the discharged metals are easily transported through groundwater, causing contamination of soil and rivers [8], as well as the degradation and death of vegetation, animals, and even direct damage to humans [9,10]. Therefore, in our research group, we are convinced that the recycling of metals that we use in our daily lives helps conserve the natural riches of the environment and favors the reduction of environmental pollution. Among metal scrap, door keys are among the most common forms of consumer waste. The door key used as a doping precursor was taken from a metal scrap yard; hence, it had advanced metal wear. Due to this, its origin of manufacture is unknown. However, it is known that most door keys are alloys of steel or brass.

Among the metals that compose these alloys, nickel, copper, and zinc are chemical elements widely used to dope TiO<sup>2</sup> nanoparticles [11–13]. Such is the case of a study carried out by Raguram, T. et al. [11], who synthesized TiO2-Ni nanoparticles using the sol–gel method. The results reported by the researchers indicate that the doping of Ni2+ in TiO<sup>2</sup> benefited the absorption of visible light, reaching a maximum of 61.04% in the degradation of the methylene blue dye in an aqueous medium. On the other hand, Hemraj, Y. et al. [12] synthesized TiO<sup>2</sup> nanoparticles doped with different Cu2+ contents (0 to 3.0% mol) using the sol–gel method for semiconductor synthesis. The results show that the TiO2-Cu nanoparticle (3.0%) has a photocatalytic efficiency superior to pure TiO2. They mentioned that the improvement in the photocatalytic performance for the photodegradation of methyl orange was due to the redshift of the bandgap energy (Eg) and the decrease in the recombination rate of the electron–hole pair of TiO2-Cu nanoparticles. In general, it has been mentioned that noble metals such as Ni and Cu are good candidates to act as traps for photogenerated electrons because the Fermi level of these ions is lower than that of TiO2. Regarding Zn use in TiO<sup>2</sup> doping, Shao, M. et al. [13] concluded that adding Zn to TiO<sup>2</sup> reduces Eg and increases optical absorption in the visible region, reaching a maximum rate of 88.14% in degradation of tetracycline. In addition to the above, many researchers have focused on multi-doping metals (bimetallic or trimetallic) over the last few years. They mention that the second metal can alter the electronic properties and the formation of active surface structures, increasing the photocatalytic activity of TiO<sup>2</sup> concerning monometallic doping [14–17]. The previous work provides evidence of the feasibility of using transition metals as monometallic or bimetallic dopants in TiO<sup>2</sup> nanoparticles using chemical reagents as doping precursors. However, we did not find any scientific report investigating the use of door-key waste for TiO<sup>2</sup> multidoping. As an additional motivation, it is important to mention that innovative strategies are currently being developed to increase the efficiency of contaminant removal and transformation [18–22]. In this context, an innovative synthesis route was designed to obtain TiO<sup>2</sup> doped with metallic ions, where the use of chemical reagents is avoided, and it was proposed to use a metallic residue (door key) as a doping precursor to improving the photocatalytic efficiency of TiO<sup>2</sup> in the mineralization of an emerging contaminant.

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

### *2.1. Chemicals and Reagents*

To synthesize TiO2, reagent grade chemicals were used without further purification. Titanium (IV) butoxide (97%), 1-butanol anhydrous (99.8%), and nitric acid (≥65%) were purchased from Sigma-Aldrich. Deionized water was used for the hydrolysis step during

the sol–gel synthesis process. A door key recovered from a metal scrapyard was used as a doping precursor.

### *2.2. Synthesis of Photocatalysts*

To analyze the effect of metal ions that make up the door key on the photocatalytic activity of TiO2, 3 samples were prepared with different percentages of metallic waste (0, 0.5, and 1.0 wt.%). The materials nomenclature is defined as T, T-DK (0.5), and T-DL (1.0) as well as the abbreviation T (TiO2) + the metal residue percentage number (DK).

The synthesis of the photocatalysts was carried out using a non-traditional sol–gel synthesis route. The waste door key (0.5 or 1.0 wt.%) for the TiO<sup>2</sup> doping was dissolved at room temperature in 1.5 mL of HNO3. *Solution A*. To synthesize TiO2, doped titanium (IV) butoxide and anhydrous 1-butanol were mixed. *Solution B*. After stirring for 30 min at room temperature, *solution A* was added drop by drop to *solution B*, continuing to stir for 30 min. To begin the hydrolysis reaction in the *A–B* mixture, 10 mL of deionized H2O was added per drop. The gel obtained was aged at ambient temperature for 24 h. Xerogel was subsequently obtained by filtering the gel with a vacuum pump to remove the solvents and then drying for 12 h at 60 ◦C. Finally, the xerogel was heat-treated at 400 ◦C for 4 h using a 4 ◦C min−<sup>1</sup> heating ramp with a static air atmosphere. Pure TiO<sup>2</sup> was synthesized using a similar methodology as described earlier for comparison purposes. The only difference is that for pure TiO<sup>2</sup> synthesis, *solution A* had no metallic ions.

### *2.3. Characterization of Photocatalysts*

The surface area of the photocatalysts was determined using the BET method. Their N<sup>2</sup> adsorption–desorption isotherms were obtained at −196 ◦C using Asap 2020 equipment after degassing samples at 250 ◦C for 12 h in a high vacuum.

The crystalline phase for each sample was determined using a Siemens D500 X-ray diffractometer (XRD) with Cu K<sup>α</sup> radiation (λ = 0.15418 nm). The diffraction patterns were obtained within the range of 10 to 80◦ at an acquisition rate of 0.02 s−<sup>1</sup> and 1 s per point. The average crystal size was calculated using the Scherrer equation.

$$\Phi = \frac{\text{K}\lambda}{\text{ $\beta$ } \cos \theta} \tag{1}$$

where Φ is the size of the crystal, K (0.91) is the form factor, λ is the wavelength of the X-rays, β is the width at half height (FWHM) of the main intensity peak, and θ is the Bragg angle.

The optical properties of the materials were obtained with diffuse reflection spectroscopy (DRS) using a Perkin-Elmer UV-Vis spectrophotometer with an integrating sphere. Absorption spectra were recorded in the range from 300 to 800 nm. A barium sulfate plate (BaSO4) was used as a reference. Based on the absorption spectra, the bandgap energies of the photocatalysts were determined using the Kubelka–Munk theory and Tauc plot [23].

The morphological characterization of composites was conducted with transmission electron microscopy (HRTEM) using a JEOEL electronic microscope model JEM2100.

The semi-quantitative elementary composition of the materials was determined with energy-dispersive X-ray spectroscopy (EDX) using an electron scanning microscope with Rontec Xflash detector model Hitachi S-4700 Type IIc. In addition, the chemical composition of the catalysts was determined using inductively coupled plasma optical emission spectroscopy (ICP-OES) Varian 725-ES, after acidic digestion (Nitric Acid) of the materials.

The XPS spectra were obtained using X-ray photoelectron spectroscopy (XPS) with Thermo Scientific apparatus equipped with a non-monochromated Mg anode as an X-ray source operated at 12 kV under vacuum (2 <sup>×</sup> <sup>10</sup>−<sup>7</sup> mbar). The binding energies were calibrated at 284.6 eV with respect to the C1s peak of the carbon samples. The XPS spectrums were deconvoluted using the PeakFit software. The peaks were fitted to a linear background and a combination of Gaussian/Lorentzian functions.

### *2.4. Evaluation of Photocatalytic Activity*

Diclofenac was used as a contaminant to evaluate the photocatalytic activity of photocatalysts. The photocatalytic degradation of 10 ppm of diclofenac (Sigma-Aldrich, Burlington, MA, USA, ≥98.5%) was carried out in a glass reactor with a volume of 0.5 L and photocatalyst load of 1.0 gL−<sup>1</sup> . The reactor was placed in a SUNTEST XLS+ solar simulator, equipped with a xenon arc lamp model XL-S-750 and a 320 nm cut-off filter. During the experiments, the temperature of the reaction system was maintained in the range of 25 to 35 ◦C. The suspension was stirred in the dark for 30 min while the air was bubbled into the system. The light was then turned on, and simultaneously the system continued to be supplied with air. To monitor the mineralization of diclofenac, samples were collected from the reactor every 30 min to up 180 min. Photolysis experiments were also carried out for comparative purposes. The progression of the mineralization of the contaminant was followed by the measurement of total organic carbon using a Shimadzu TOC-VSCH analyzer. The following equation determined the mineralization of diclofenac:

$$\% \text{TOC} = \frac{\text{TOC}\_0 - \text{TOC}\_f}{\text{TOC}\_0} \times 100 \tag{2}$$

where *TOC*<sup>0</sup> is the mineralization at time zero and *TOC<sup>f</sup>* is the final mineralization at each instant.

### **3. Results**

### *3.1. Sample Characterization*

### 3.1.1. XRD

The XRD patterns for pure and doped samples are illustrated in Figure 1. In all diffractograms, peaks are observed at angles 25.33, 37.8, 48.06, 53.9, and 55.08, characteristic of TiO2-anatase (JCPDS card no. 21-1272). The crystalline phase obtained corresponds to that reported by other investigators. Other studies mention that the crystalline phase of anatase occurs when heat treatment varies between 200 and 500 ◦C. It was said that the amount and physicochemical properties (high ionic radii and electronic status) of the doping metal can influence the rate of transformation of anatase to the rutile phase [24,25]. Such is the case of a study conducted by Hampel et al. [25]; they observed that TiO<sup>2</sup> doped with 5 wt.% Cu ions, calcined at 450 ◦C, contained an anatase phase and a rutile phase fraction, which increased with increasing Cu concentration from 5 to 10 wt.%. In our case, the dopant concentration does not influence the anatase–rutile transition because the maximum theoretical concentration of doping ions in TiO<sup>2</sup> is less than 1 wt.%. *Water* **2022**, *14*, x FOR PEER REVIEW 5 of 15

> It is convenient to have TiO2 nanoparticles with an anatase crystalline structure for this research. There is evidence that other polymorphic forms of this material are less

> No characteristic peaks were associated with doping Cu, Zn, or Ni (metal oxide). It is known that X-ray diffractometers are not sensitive to levels of impurities lower than 5%. For this reason, the absence of characteristic peaks of the dopant ions can be attributed to the fact that the percentages of door lock used during the synthesis of the doped TiO2 were less than 1 wt.%. Likewise, this also suggests that the doping ions could be highly

> For doped TiO2 nanoparticles, the maximum intensity of the anatase phase crystalline plane (101) decreases. This indicates that the crystallization of TiO2 anatase is restricted due to the increase in the content of the dopant (0.5 to 1.0 wt.%). As shown in Table 1, the increase in the doping content leads to an average particle size smaller than the pure material. The decrease in crystalline nature has been reported by other authors who used Cu and Zn salts as doping precursors. They mentioned that a mechanism of fixation of doping ions at the grain boundary inhibits anatase nanoparticle growth [29,30,31].

> > **Zn wt.%**

**Sol. A 1 EDX ICP-OES Sol. A 1 EDX ICP-OES DRX SEM** T nd nd nd nd nd nd 19.50 21 3.23

T-DK (0.5) 0.21 0.19 0.23 0.13 0.12 0.11 15.7 16 2.90 ±0.009 ±0.002 ±0.014 ±0.005 ±0.004 ±0.003

T-DK (1.0) 0.47 0.43 0.44 0.24 0.22 0.21 11.3 11 2.76 ±0.011 ±0.015 ±0.009 ±0.008 ±0.007 ±0.006

**Average Particle Size (nm)**

**Eg (eV)**

reduced to superoxide radicals (O2•−) by photoexciting electrons from BV to BC, thus functioning as an electron trap, inhibiting the recombination of e−/h+ pairs. This increases the half-life of the redox reactions that benefit the formation of hydroxyl radicals (OH•) with sufficient oxidation potential for the photocatalytic degradation of recalcitrant compounds. Along with the above, the redox potential of BV in anatase is more negative, making it more competitive than the rutile phase for oxidation reactions [26,27,28].

**Figure 1.** X-ray diffraction (XRD) patterns of TiO2 and doped-TiO2. **Figure 1.** X-ray diffraction (XRD) patterns of TiO<sup>2</sup> and doped-TiO<sup>2</sup> .

**Table 1.** Material characterization results of TiO2 and doped-TiO2.

1 Sol. A: Solution resulting from the digestion of the dopant precursor.

**Cu wt.%**

3.1.2. HRTEM and Elemental Analysis

dispersed in the titania support.

**Catalyst**

It is convenient to have TiO<sup>2</sup> nanoparticles with an anatase crystalline structure for this research. There is evidence that other polymorphic forms of this material are less effective in photocatalytic reactions. This is due to the higher adsorption capability of O<sup>2</sup> on the surface of the anatase phase than in the rutile phase. The adsorbed oxygen is reduced to superoxide radicals (O<sup>2</sup> •−) by photoexciting electrons from BV to BC, thus functioning as an electron trap, inhibiting the recombination of e−/h<sup>+</sup> pairs. This increases the half-life of the redox reactions that benefit the formation of hydroxyl radicals (OH• ) with sufficient oxidation potential for the photocatalytic degradation of recalcitrant compounds. Along with the above, the redox potential of BV in anatase is more negative, making it more competitive than the rutile phase for oxidation reactions [26–28].

No characteristic peaks were associated with doping Cu, Zn, or Ni (metal oxide). It is known that X-ray diffractometers are not sensitive to levels of impurities lower than 5%. For this reason, the absence of characteristic peaks of the dopant ions can be attributed to the fact that the percentages of door lock used during the synthesis of the doped TiO<sup>2</sup> were less than 1 wt.%. Likewise, this also suggests that the doping ions could be highly dispersed in the titania support.

For doped TiO<sup>2</sup> nanoparticles, the maximum intensity of the anatase phase crystalline plane (101) decreases. This indicates that the crystallization of TiO<sup>2</sup> anatase is restricted due to the increase in the content of the dopant (0.5 to 1.0 wt.%). As shown in Table 1, the increase in the doping content leads to an average particle size smaller than the pure material. The decrease in crystalline nature has been reported by other authors who used Cu and Zn salts as doping precursors. They mentioned that a mechanism of fixation of doping ions at the grain boundary inhibits anatase nanoparticle growth [29–31].

**Table 1.** Material characterization results of TiO<sup>2</sup> and doped-TiO2.


<sup>1</sup> Sol. A: Solution resulting from the digestion of the dopant precursor.

### 3.1.2. HRTEM and Elemental Analysis

In the HRTEM micrographs in Figure 2, pure and doped TiO<sup>2</sup> nanoparticles exhibit spherical morphology. After doping, it is also observed that the average crystal size decreases compared to pure TiO2. This phenomenon has already been explained in the XRD diffractogram discussion. Based on the histograms in Figure 2, for T and T-DK (1.0), the average particle size was similar to those obtained with XRD.

Although it has been shown that the surface area is not decisive in the photocatalytic process, as a result of particle size decrease, the specific surface area increased gradually from 58.8 <sup>±</sup> 0.8 m2<sup>g</sup> −1 for T, 76 <sup>±</sup> 0.4 m2<sup>g</sup> −1 for T-DK (0.5), and 88.7 <sup>±</sup> 0.5 m2<sup>g</sup> −1 for T-DK (1.0). Increased dopants in TiO<sup>2</sup> inhibit particle growth; consequently, the narrow pore size distribution and the surface becomes larger [32].

From the HRTEM micrography in Figure 2, it is difficult to determine the presence of doping ions. Nevertheless, the elementary EDS and ICP-OES analysis results, summarized in Table 1, show that doped TiO<sup>2</sup> nanoparticles contain Cu and Zn doping ions. According to the EDS images in Figure 2, these dopants are homogeneously distributed over the surface of the TiO2. The presence of Ni ions was not detected. As the dopant precursor was obtained from a junkyard, it had advanced metal wear, so there was no longer any nickel coating that this type of door key usually has. So, the absence of nickel in the elemental

analysis performed by EDS and ICP-OES was an expected result. This means that only copper and zinc were considered doping ions. XRD diffractogram discussion. Based on the histograms in Figure 2, for T and T-DK (1.0), the average particle size was similar to those obtained with XRD.

In the HRTEM micrographs in Figure 2, pure and doped TiO2 nanoparticles exhibit spherical morphology. After doping, it is also observed that the average crystal size decreases compared to pure TiO2. This phenomenon has already been explained in the

*Water* **2022**, *14*, x FOR PEER REVIEW 6 of 15

**Figure 2.** HRTEM micrographs, histograms, and EDS spectra of TiO2 pure and T-DK (1.0). **Figure 2.** HRTEM micrographs, histograms, and EDS spectra of TiO<sup>2</sup> pure and T-DK (1.0).

Regarding the weight percentage of Cu and Zn in TiO2, EDS and ICP-OES (Table 1) show differences between the theoretical weight percentage of Cu and Zn concerning the real percentage. To explain this difference, elemental analysis by ICP-OES was performed

on solution A (solution resulting from the digestion of the metal residue). The results show that during the digestion carried out in an open system, part of the initial concentration of the doping ions is lost (Table 1). The efficacy of the digestion method, among other variables, will be investigated in future projects. on solution A (solution resulting from the digestion of the metal residue). The results show that during the digestion carried out in an open system, part of the initial concentration of the doping ions is lost (Table 1). The efficacy of the digestion method, among other variables, will be investigated in future projects.

This means that only copper and zinc were considered doping ions.

Although it has been shown that the surface area is not decisive in the photocatalytic process, as a result of particle size decrease, the specific surface area increased gradually from 58.8 ± 0.8 m2g−1 for T, 76 ± 0.4 m2g−1 for T-DK (0.5), and 88.7 ± 0.5 m2g−1 for T-DK (1.0). Increased dopants in TiO2 inhibit particle growth; consequently, the narrow pore size

From the HRTEM micrography in Figure 2, it is difficult to determine the presence of doping ions. Nevertheless, the elementary EDS and ICP-OES analysis results, summarized in Table 1, show that doped TiO2 nanoparticles contain Cu and Zn doping ions. According to the EDS images in Figure 2, these dopants are homogeneously distributed over the surface of the TiO2. The presence of Ni ions was not detected. As the dopant precursor was obtained from a junkyard, it had advanced metal wear, so there was no longer any nickel coating that this type of door key usually has. So, the absence of nickel in the elemental analysis performed by EDS and ICP-OES was an expected result.

Regarding the weight percentage of Cu and Zn in TiO2, EDS and ICP-OES (Table 1) show differences between the theoretical weight percentage of Cu and Zn concerning the real percentage. To explain this difference, elemental analysis by ICP-OES was performed

*Water* **2022**, *14*, x FOR PEER REVIEW 7 of 15

distribution and the surface becomes larger [32].

#### 3.1.3. DRS 3.1.3. DRS

The optical properties of these materials were obtained using DRS. The absorption spectra are shown in Figure 3. Bandgap values (Eg) are shown in Table 1. Pure TiO<sup>2</sup> exhibits strong absorption at 396 nm (3.23 eV), assigned to the charge transfer of the metal-ligand in Ti4+(3d)-O<sup>2</sup> -(2p). For T-DK (0.5) and T-DK (1.0), the door key used as a dopant influenced the optical properties of TiO<sup>2</sup> since the optical absorption band decreased below 396 nm with the increase in the percentage by weight of dopant ions. The T-DK (0.5) material presented a bandgap of 2.90 eV, and the T-DK (1.0) material showed a bandgap of 2.76 eV. It is known that the value of the bandgap of ZnO (Eg ≈ 3.2 eV) is similar to that of TiO<sup>2</sup> (Eg ≈ 3.23 eV) [33], so zinc does not promote the decrease in Eg. The decrease is due to a redistribution of the electric charge of TiO<sup>2</sup> caused by copper ions in interstitial positions or to copper oxides (Eg ≈ 2.5 eV) formed [34]. The obtained Eg values indicate that in our materials, the valence electrons can be transferred to the conduction band using lower energy (visible light) than that required to activate pure TiO<sup>2</sup> (UV light). The optical properties of these materials were obtained using DRS. The absorption spectra are shown in Figure 3. Bandgap values (Eg) are shown in Table 1. Pure TiO2 exhibits strong absorption at 396 nm (3.23 eV), assigned to the charge transfer of the metalligand in Ti4+(3d)-O2-(2p). For T-DK (0.5) and T-DK (1.0), the door key used as a dopant influenced the optical properties of TiO2 since the optical absorption band decreased below 396 nm with the increase in the percentage by weight of dopant ions. The T-DK (0.5) material presented a bandgap of 2.90 eV, and the T-DK (1.0) material showed a bandgap of 2.76 eV. It is known that the value of the bandgap of ZnO (Eg ≈ 3.2 eV) is similar to that of TiO2 (Eg ≈ 3.23 eV) [33], so zinc does not promote the decrease in Eg. The decrease is due to a redistribution of the electric charge of TiO2 caused by copper ions in interstitial positions or to copper oxides (Eg ≈ 2.5 eV) formed [34]. The obtained Eg values indicate that in our materials, the valence electrons can be transferred to the conduction band using lower energy (visible light) than that required to activate pure TiO2 (UV light).

**Figure 3.** Absorption spectrum of T, T-DK (0.5), and T-DK (1.0). Insets: bandgap energy values: linear part of the plot extrapolated to the X-axis.

For T-DK (1.0), the absorption band at 600 nm suggests a <sup>2</sup>B1g→2B2g characteristic of octahedral coordination with tetragonal distortion around Cu2+ [34,35]. For low concentrations of copper, T-DK (0.5), no characteristic bands corresponding to metallic ion species were observed. Likewise, no interfacial charge transfer bands were observed from the valence band of TiO<sup>2</sup> to the valence band of zinc oxides. This is probably due to the low weight concentration of the dopant ions or their good dispersion on the titania surface.

### 3.1.4. X-ray Photoelectron Spectroscopy

The XPS spectra of T and T-DK (1.0) were obtained to determine the oxidation state of the material components. The oxidation states in the O 1s, Cu2p, and Zn 2p regions were obtained by deconvoluting their peaks. The PeakFit software version 4.1.2, AISN Software Inc, was used to this effect. The correlation coefficients (r<sup>2</sup> ) of the deconvoluted peaks were higher than 0.99. The spectrum in Figure 4a shows bands at 458.5 and 464.1 eV, characteristic of the binding energy of Ti 2p3/2 and Ti 2p1/2 in TiO<sup>2</sup> [36]. The observed band comprises two peaks in the O1s region of pure TiO2. One of these, located at 530.02 eV, is assigned to ionic oxygen in the crystalline array (O-Ti4+). The small peak at 531.82 eV is

related to adsorbed OH groups, chemisorbed O species, or oxygen vacancies [37]. These same bands were observed in the Ti 2p (Figure 4c) and O 1s (Figure 4d) regions of the XPS spectra obtained for T-DK (1.0). However, a shift in the binding energies of the Ti 2p and O 1s bands is observed for this material. It is known that the binding energy depends on the oxidation state and the local chemical environment of titanium and oxygen [38,39]. Since the electronegativity of Cu (1.9) and Zn (1.65) is greater than that of Ti (1.54), the electron density around the oxygen ions decreases, which causes an increase in the binding energy. *Water* **2022**, *14*, x FOR PEER REVIEW 9 of 15

**Figure 4.** XPS spectra of TiO2 and T-DK (1.0): (**a**) and **b**) Ti 2p, (**c**) and (**d**) O 1s, (**e**) Cu 4d, and (**f**) Zn **Figure 4.** XPS spectra of TiO<sup>2</sup> and T-DK (1.0): (**a**,**b**) Ti 2p, (**c**,**d**) O 1s, (**e**) Cu 4d, and (**f**) Zn 4d.

4d. To develop the above, the Cu 2p (Figure 4e) and Zn 2p (Figure 4f) XPS spectra from the T-DL (1.0) photocatalyst were analyzed. In Figure 4e, the binding energies at ~933.78 eV correspond to Cu+ from Cu2O, while the binding energies ~935.76 eV can be assigned to Cu2+ in the form of CuO or Cu(OH)2 [25,39,40]. In the case of zinc ions, a typical peak of Zn 2p3/2 is observed in Figure 4f. The deconvolution of this peak produced two peaks at 1021.3 and 1023.8 eV. The strongest peak located at 1021.3 eV is associated with Zn2+ ions in the ZnO with an arrangement of wurtzite crystal. Peak at 1023.8 eV belongs to Zn2+ ions To develop the above, the Cu 2p (Figure 4e) and Zn 2p (Figure 4f) XPS spectra from the T-DL (1.0) photocatalyst were analyzed. In Figure 4e, the binding energies at ~933.78 eV correspond to Cu<sup>+</sup> from Cu2O, while the binding energies ~935.76 eV can be assigned to Cu2+ in the form of CuO or Cu(OH)<sup>2</sup> [25,39,40]. In the case of zinc ions, a typical peak of Zn 2p3/2 is observed in Figure 4f. The deconvolution of this peak produced two peaks at 1021.3 and 1023.8 eV. The strongest peak located at 1021.3 eV is associated with Zn2+ ions in the ZnO with an arrangement of wurtzite crystal. Peak at 1023.8 eV belongs to Zn2+ ions in Zn(OH)<sup>2</sup> [41–43].

in Zn(OH)2 [41,42,43]. The ionic radius of Ti4+, Cu2+/Cu+, and Zn2+ is 0.061, 0.73/0.77, and 0.74 Å, respectively The ionic radius of Ti4+, Cu2+/Cu<sup>+</sup> , and Zn2+ is 0.061, 0.73/0.77, and 0.74 Å, respectively [44]. These data indicate a significant difference between the values of the ionic radii

[44]. These data indicate a significant difference between the values of the ionic radii of the dopants with titanium. As a result, the substitution process of Ti ions by Cu2+/Cu+ and

between them is less than 20%. Theoretical studies for copper ions argue that, for the substitution of titanium ions in the crystalline structure, the maximum concentration of copper should be 0.3% [39]. As a result, we can report that some copper and zinc ions were aggregated as oxides on the surface of TiO2 nanoparticles, creating a heterojunction

of the dopants with titanium. As a result, the substitution process of Ti ions by Cu2+/Cu<sup>+</sup> and Zn2+ is limited by the difference between their ionic radii. According to the principles of Hume-Rothery [45], lattice substitution between atoms can only happen if the difference between them is less than 20%. Theoretical studies for copper ions argue that, for the substitution of titanium ions in the crystalline structure, the maximum concentration of copper should be 0.3% [39]. As a result, we can report that some copper and zinc ions were aggregated as oxides on the surface of TiO<sup>2</sup> nanoparticles, creating a heterojunction between these semiconductor materials. Other doping ions held interstitial positions within the crystalline network of the TiO<sup>2</sup> anatase.

According to the results of the material characterization and the references used for their discussion, the interaction between the support and the copper and zinc ions is similar to that reported by other investigations in which the dopant precursor is a chemical reagent.

### *3.2. Photocatalytic Activity Measurement*

To evaluate the photocatalytic activity of the synthesized materials, a solution of diclofenac at a concentration of 10 ppm was used. Total organic carbon (TOC) values are summarized in Table 2. All the materials showed photocatalytic activity in the degradation of diclofenac (Figure 5a,b). Nevertheless, for doped materials, the increased content of doping ions enhanced the photocatalytic activity of TiO2. The best activity was achieved when T-DK (1.0) was used as a photocatalyst, obtaining up to 94% of mineralized diclofenac. The values for the rate constants (Table 2 and Figure 5b) are consistent with what is seen in Figure 5a. This means that with T-DK (1.0), the maximum mineralization is achieved, but this transformation is also carried out in less time than with T and T-DK (0.5).

**Figure 5.** (**a**) Total organic carbon, (**b**) kinetic constants, (**c**) PL spectra of undoped and codoped TiO<sup>2</sup> .


**Table 2.** Apparent first-order rate constants obtained for the diclofenac mineralization.

The correlation of material characterization results with photocatalytic test results suggests that the increase in photocatalytic activity is due to two factors:

i. First, the doping ions induced the reduction of Eg in TiO<sup>2</sup> (see Table 1). The decrease in the value of Eg for T-DK (0.5) and T-DK (1.0) indicates that these materials can be activated with visible light radiation. This means solar light absorption is more efficient in doped TiO<sup>2</sup> than pure TiO<sup>2</sup> [46]. This increases the efficiency of generating electron–hole pairs that initiate redox reactions that directly or indirectly produce the hydroxyl radicals that cause the pollutant to be mineralized. Due to the many possible reaction mechanisms during the diclofenac mineralization process, those considered the main ones in the photocatalytic mechanism are given below [24].

$$\text{Catalyst} + hv\_{(Sunlight)} \rightarrow \text{Catalyst} \left(e\_{CB}^{-} + h\_{VB}^{+} \right)$$

$$h^{+} + H\_{2}O \rightarrow HO^{\bullet} + H^{+} $$

$$h^{+} + OH^{-} \rightarrow HO^{\bullet}$$

Hydroxyl radical attack:

$$D\text{CF} + HO^\bullet \rightarrow \text{Degradiation products}$$

Oxidation by the positive hole:

$$D\text{CF} + h^+ \rightarrow \text{Oxidation products}$$

ii. Second, the synergistic effect of the dopant species inhibited the recombination of the e−/h<sup>+</sup> pairs. This information was obtained by analyzing the charge carrier recombination of each synthesized material. In the emission spectra in Figure 5c, pure TiO<sup>2</sup> obtained the higher intensity emission spectra, meaning rapid recombination of the electron–hole pairs. Contrary to TiO2, the intensity of the emission spectra was lower when the percentage by weight of the doping ions increased, which suggests a low recombination rate for the electron–hole pairs photogenerated in T-DK (0.5) and T-DK (1.0). According to the results obtained using XPS, in doped TiO<sup>2</sup> Cu2+ and Zn2+ ions in interstitial positions and oxides of the doping ions coexist. The low recombination of photogenerated charge carriers at T-DK (0.5) and T-DK (1.0) can be understood due to the positions of the band edges of the oxides in the heterojunction. The measured conduction band (CB) potential values of TiO<sup>2</sup> and CuO are –0.35 and +0.12 V (vs. SCE), respectively. The valence band of TiO<sup>2</sup> is lower than ZnO by about 0.36 V (vs. NHE), and this is superior to CuO by approximately 0.20 V (vs. SHE). The relative position difference of the energy band of CuO and ZnO charge transfer occurs between them and TiO2. Thus, an electron photogenerated in TiO<sup>2</sup> is transferred from the conduction band of this semiconductor to ZnO and CuO, acting as an electron trap to inhibit their recombination. Concurrent with the above, hole transfer can arise from the valence band (VB) of TiO<sup>2</sup> to the VB of ZnO and CuO [43,44,47,48]. Therefore, in addition to the decrease in Eg, the coupled effect between energy bands of TiO2, CuO, and ZnO was an essential factor to suppress the recombination of the electron–hole pairs, improving the photocatalytic activity of doped TiO2.

An important characteristic of photocatalysts is their chemical stability after several recycling cycles. Therefore, the reuse of T-DK (1.0) was evaluated during five consecutive reuse cycles. The percentage results in Figure 6a show that during the five cycles of reuse, the mineralization efficiency of diclofenac was in the range of 92.5 to 95%, with a standard deviation of ±1.48. Consequently, no significant change in the mineralization efficiency of diclofenac was observed. The stability in the photocatalytic activity is because the concentration of the dopant ions (Figure 6a) and the crystal structure of the TiO2-anatase (Figure 6b) does not change before or after the reuse cycles. A similar effect occurs for the photocatalyst morphology determined by SEM in Figure 6c,d. This means that the photocatalyst T-DK (1.0) has a high activity and stability during the five cycles of diclofenac photocatalytic mineralization. *Water* **2022**, *14*, x FOR PEER REVIEW 12 of 15 deviation of ±1.48. Consequently, no significant change in the mineralization efficiency of diclofenac was observed. The stability in the photocatalytic activity is because the concentration of the dopant ions (Figure 6a) and the crystal structure of the TiO2-anatase (Figure 6b) does not change before or after the reuse cycles. A similar effect occurs for the photocatalyst morphology determined by SEM in Figure 6c,d. This means that the photocatalyst T-DK (1.0) has a high activity and stability during the five cycles of diclofenac photocatalytic mineralization.

**Figure 6.** Reuse cycles using T-DK (1.0) and elemental analysis of Cu and Zn before and after the reuse cycle (**a**). DRX diffractograms (**b**) and SEM micrographs of T-DK (1.0) before and after five cycles of reuse (**c**,**d**). **Figure 6.** Reuse cycles using T-DK (1.0) and elemental analysis of Cu and Zn before and after the reuse cycle (**a**). DRX diffractograms (**b**) and SEM micrographs of T-DK (1.0) before and after five cycles of reuse (**c**,**d**).

#### **4. Conclusions 4. Conclusions**

This work shows that incorporating Cu and Zn ions into TiO2 is possible when a door key is used as a doping precursor. Material characterization showed that these ions were integrated into the support as Cu+/Cu2+ and Zn2+, occupying interstitial positions or forming heterojunctions between copper and zinc oxides with titania. The modification with both ions had a dual effect dependent on the concentration of the doping metals. First, an inverse correlation was observed between the concentration of doping ions and the Eg. This resulted in greater light absorption within the visible range. Second, it was shown that the impurities in TiO2 acted as an electron trap, so the recombination of the e- /h+ pairs was lower for T-DK (1.0) than for TiO2. Therefore, the photocatalytic activity of the synthesized materials occurred in the following order: T < T-DK (0.5) < T-DK (1.0). This work shows that incorporating Cu and Zn ions into TiO<sup>2</sup> is possible when a door key is used as a doping precursor. Material characterization showed that these ions were integrated into the support as Cu+/Cu2+ and Zn2+, occupying interstitial positions or forming heterojunctions between copper and zinc oxides with titania. The modification with both ions had a dual effect dependent on the concentration of the doping metals. First, an inverse correlation was observed between the concentration of doping ions and the Eg. This resulted in greater light absorption within the visible range. Second, it was shown that the impurities in TiO<sup>2</sup> acted as an electron trap, so the recombination of the e-/h<sup>+</sup> pairs was lower for T-DK (1.0) than for TiO2. Therefore, the photocatalytic activity of the synthesized materials occurred in the following order: T < T-DK (0.5) < T-DK (1.0).

The results obtained in this work confirm previous state-of-the-art findings: Doping TiO2 with Cu and Zn ions or other metal ions is an efficient strategy to improve the photocatalytic activity of titania. As an additional contribution, it is possible to use The results obtained in this work confirm previous state-of-the-art findings: Doping TiO<sup>2</sup> with Cu and Zn ions or other metal ions is an efficient strategy to improve the photo-

discarded door keys as a doping precursor to improving the photocatalytic activity of TiO2. Finally, these results open the door for future research. Other variables can be catalytic activity of titania. As an additional contribution, it is possible to use discarded door keys as a doping precursor to improving the photocatalytic activity of TiO2. Finally, these results open the door for future research. Other variables can be considered: method of digestion of the metal residue, method of synthesis of materials, pH of synthesis, and calcination temperatures, among other variables. These variables were not considered in this first study, as the main objective was to investigate the feasibility of using door keys as doping precursors.

**Author Contributions:** Conceptualization, Visualization, Formal analysis, Investigation, Validation, Writing—original draft, Writing—review and editing, D.E.J.-C.; Writing—original draft, Writing review and editing, Resources, J.G.T.-T.; Writing—original draft, Writing—review and editing, Resources, A.H.-R.; Writing—original draft, Writing—review and editing, J.C.A.-P.; Writing—original draft, Writing—review and editing, A.C.-U.; Writing—original draft, Writing—review and editing, A.E.E.d.l.M.; Writing—original draft, Writing—review and editing, S.G.; Writing—original draft, Writing—review and editing, A.A.S.-P.; Conceptualization, Formal analysis, Funding acquisition, Investigation, Methodology, Project administration, Resources, Supervision, Validation, Visualization, Writing—original draft, Writing—review and editing, A.C.-G. 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.

**Data Availability Statement:** Data are contained within the article.

**Acknowledgments:** This research was supported by the National Council of Science and Technology (CONACYT) through the scholarship granted to Dany Edgar Juárez Cortazar.

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

### **References**


## *Review* **Recent Advances of Emerging Organic Pollutants Degradation in Environment by Non-Thermal Plasma Technology: A Review**

**Yongjian He, Wenjiao Sang \*, Wei Lu, Wenbin Zhang, Cheng Zhan and Danni Jia**

School of Civil Engineering and Architecture, Wuhan University of Technology, Wuhan 430070, China; hyjwhut@126.com (Y.H.); wluemail@163.com (W.L.); zhangwenbin1114@126.com (W.Z.); wlzhancheng@126.com (C.Z.); jdnwz9998@163.com (D.J.)

**\*** Correspondence: whlgdxswj@126.com

**Abstract:** Emerging organic pollutants (EOPs), including endocrine disrupting compounds (EDCs), pharmaceuticals and personal care products (PPCPs), and persistent organic pollutants (POPs), constitute a problem in the environmental field as they are difficult to completely degrade by conventional treatment methods. Non-thermal plasma technology is a novel advanced oxidation process, which combines the effects of free radical oxidation, ozone oxidation, ultraviolet radiation, shockwave, etc. This paper summarized and discussed the research progress of non-thermal plasma remediation of EOPs-contaminated water and soil. In addition, the reactive species in the process of non-thermal plasma degradation of EOPs were summarized, and the degradation pathways and degradation mechanisms of EOPs were evaluated of selected EOPs for different study cases. At the same time, the effect of non-thermal plasma in synergy with other techniques on the degradation of EOPs in the environment was evaluated. Finally, the bottleneck problems of non-thermal plasma technology are summarized, and some suggestions for the future development of non-thermal plasma technology in the environmental remediation were presented. This review contributes to our better understanding of non-thermal plasma technology for remediation of EOPs-contaminated water and soil, hoping to provide reference for relevant practitioners.

**Keywords:** advanced oxidation processes; discharge plasma; reactive species; environmental remediation; combination system; degradation mechanism

### **1. Introduction**

Over the past few decades, increasing industrial, agricultural, and human activities have promoted the use of chemicals [1,2]. In addition, the development of human society and the wrong understanding of personal safety have increased the chemical load in water and soil environment. In many developing countries, untreated sewage is also used for agricultural purposes, resulting in substandard or untreated wastewater that adds many pollutants to the food chain [3,4]. Therefore, appropriate technologies are needed to eliminate pollutants in water and soil environment for the sake of protection for both human health and the environment.

Emerging organic pollutants (EOPs) are a kind of organic pollutants which have no environmental monitoring standards or emission standards and have negative effects on ecology and human health [5]. EOPs include endocrine disrupting compounds (EDCs), pharmaceutical and personal care products (PPCPs), and persistent organic pollutants (POPs) [6,7]. EOPs may be candidates for future regulation because of their potential risks to the environment and human health, the continuous entry into the environment and the fact that even the most modern wastewater treatment plants (WWTPs) cannot completely convert/remove these compounds [8]. In recent years, with the improvement of environmental analysis, these substances have been frequently detected in the environment, such as sewage, surface water, drinking water [9], and soil [10]. The continuous detection

**Citation:** He, Y.; Sang, W.; Lu, W.; Zhang, W.; Zhan, C.; Jia, D. Recent Advances of Emerging Organic Pollutants Degradation in Environment by Non-Thermal Plasma Technology: A Review. *Water* **2022**, *14*, 1351. https://doi.org/ 10.3390/w14091351

Academic Editors: Dionysios (Dion) Demetriou Dionysiou, Yujue Wang and Huijiao Wang

Received: 29 March 2022 Accepted: 19 April 2022 Published: 21 April 2022

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

**Copyright:** © 2022 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/).

of EOPs brings new challenges to environmental pollution control and makes the treatment of EOPs become an international research hotspot.

Currently, there are many methods for organic pollutants degradation in water environment: bioremediation [11,12], advanced oxidation processes (AOPs) [13–15], adsorption process [16,17], membrane treatment [18,19], and combination process [20]. Compared with water remediation, organic pollution in soil is so subtle that it is difficult to detect. In addition, soil remediation tends to be costlier and takes longer to complete. Therefore, it is urgent to develop effective and convenient remediation technology for organic contaminated soil. At present, many methods have been developed for different organic contaminated soils, including physical remediation (e.g., thermal desorption [21], soil vapor extraction [22]), chemical remediation (e.g., soil washing [23], electrochemical remediation [24], chemical oxidation remediation [25]), and biological remediation (e.g., microbial remediation [26], phytoremediation [27]). With the exception of AOPs, most of these technologies either transfer contaminants from one phase to another rather than complete degradation and mineralization, or are not efficient when the concentration of organic pollutants is at low levels. However, bioremediation often requires long treatment time and it is difficult to reduce the pollution level below the standard.

Non-thermal plasma may be a viable alternative to more common AOPs due to its comparable energy requirements for contaminant degradation and its ability to operate without any additional chemicals [28]. Non-thermal plasma technology has been used as a method to degrade EOPs, including EDCs [29,30] (e.g., pesticides, industrial chemicals, steroids), PPCPs [31–34] (e.g., antibiotics, antidepressant, anti-inflammatories, antimicrobials, surfactants), and POPs [35,36] (e.g., polychlorinated biphenyls, polycyclic aromatic hydrocarbons, organochlorine pesticides). In the review of Magureanu et al. [37], research on the degradation of various pharmaceutical compounds by non-thermal plasma was discussed, and the removal efficiency of target compounds and the energy yield of plasma technology were compared and discussed. Russo et al. [38] summarized the research works on the removal and mineralization of organic pollutants in water by the combination of non-thermal plasma and catalyst. They concluded that the catalyst played an important role in improving the performance of the plasma system. Zhang et al. [39] introduced several typical non-thermal plasma sources for remediation of organic contaminated soil. The effects of different important parameters (such as applied voltage, reactor configuration, soil properties, type of feed air, and gas flow rate) on the remediation performance were discussed. Guo et al. [40] supplemented the research on the mechanism and process of repairing organic contaminated soil by discharge plasma on the basis of previous reviews. Figure 1 shows bibliography data of papers related to non-thermal plasma technology for the degradation of pollutants published in the last decade. As shown in Figure 1a, the number of published papers using non-thermal plasma as a means of pollutant degradation in the last decade has always maintained high and is gradually increasing. Especially, PPCPs and POPs have become hotspots in this field (as shown in Figure 1b).

During the past decade, a number of studies have reported the degradation of EOPs in the environment by non-thermal plasma. At present, the review of non-thermal plasma mainly focuses on water remediation [41–45], and a few reviews have also summarized its application in soil remediation [39,40], but there is a lack of comprehensive review on water and soil remediation. Therefore, this paper focuses on the degradation of different types of EOPs (i.e., EDCs, PPCPs, and POPs) in water and soil environment by the non-thermal plasma technology. Since the reactive species in the plasma system play an indispensable role in the degradation of pollutants, the generation of reactive species in non-thermal plasma is reviewed. In addition, this review also summarizes the research progress of degradation pathways of different kinds of EOPs by non-thermal plasma. An important part of this work is devoted to the combination of non-thermal plasma with various other technologies to compensate for the shortcomings of plasmas alone. In this review, the bottleneck problems of non-thermal plasma and the future prospects are also presented.

#### During the past decade, a number of studies have reported the degradation of EOPs **2. Degradation of Emerging Organic Pollutants by Non-Thermal Plasma Technology**

in the environment by non-thermal plasma. At present, the review of non-thermal plasma mainly focuses on water remediation [41–45], and a few reviews have also summarized its application in soil remediation [39,40], but there is a lack of comprehensive review on water and soil remediation. Therefore, this paper focuses on the degradation of different types of EOPs (i.e., EDCs, PPCPs, and POPs) in water and soil environment by the nonthermal plasma technology. Since the reactive species in the plasma system play an indispensable role in the degradation of pollutants, the generation of reactive species in nonthermal plasma is reviewed. In addition, this review also summarizes the research progress of degradation pathways of different kinds of EOPs by non-thermal plasma. An important part of this work is devoted to the combination of non-thermal plasma with various other technologies to compensate for the shortcomings of plasmas alone. In this review, the bottleneck problems of non-thermal plasma and the future prospects are also presented. **2. Degradation of Emerging Organic Pollutants by Non-Thermal Plasma Technology**  Different kinds of EOPs may cause harm to human health and ecological environment; so, it is necessary to find appropriate methods to deal with EOPs. EDCs, also known as environmental hormones, can bind to hormone receptors in organisms and disrupt normal metabolism in the endocrine system [46]. Many researchers have reported their toxic effects on human health and the environment [47]. PPCPs are the most widely used chemical reagents in animal husbandry, agriculture, and human daily life. However, they have the potential to cause serious ecotoxicological problems and pose a great threat to ecosystems or organisms [48–50]. It is worth mentioning that antibiotics with a certain concentration level in the environment for a long time may not only have toxic effects on some Different kinds of EOPs may cause harm to human health and ecological environment; so, it is necessary to find appropriate methods to deal with EOPs. EDCs, also known as environmental hormones, can bind to hormone receptors in organisms and disrupt normal metabolism in the endocrine system [46]. Many researchers have reported their toxic effects on human health and the environment [47]. PPCPs are the most widely used chemical reagents in animal husbandry, agriculture, and human daily life. However, they have the potential to cause serious ecotoxicological problems and pose a great threat to ecosystems or organisms [48–50]. It is worth mentioning that antibiotics with a certain concentration level in the environment for a long time may not only have toxic effects on some sensitive organisms, but also lead to the generation, maintenance, transfer and transmission of antibiotic-resistant bacteria (ARB), and antibiotic resistance genes (ARGs) under selective pressure [51]. Since the adoption of the "Stockholm Convention" by the United Nations Environment Program (UNEP) in 2001, awareness of the potential risks of POPs in the environment and the need to remove POPs from the environment have become more urgent. POPs are generally considered to have three basic physical and chemical properties: persistence, lipophilicity, and long-distance mobility. These properties enable them to perform bioamplification and bioaccumulation in animals and seriously harm the health of humans and the natural environment. Therefore, it is imperative to develop environmentally friendly removal methods. Faced with this environmental problem, many researchers have focused on AOPs to eliminate EOPs that are resistant to conventional treatment processes [52–54]. Non-thermal plasma may be a viable alternative to more common AOPs due to its comparable energy requirements for contaminant degradation and its ability to operate without any additional chemicals. The specific research results of non-thermal plasma in water and soil remediation are introduced below.

#### sensitive organisms, but also lead to the generation, maintenance, transfer and transmis-*2.1. Water Remediation*

sion of antibiotic-resistant bacteria (ARB), and antibiotic resistance genes (ARGs) under selective pressure [51]. Since the adoption of the "Stockholm Convention" by the United Nations Environment Program (UNEP) in 2001, awareness of the potential risks of POPs in the environment and the need to remove POPs from the environment have become more urgent. POPs are generally considered to have three basic physical and chemical Researchers have paid extensive attention to the degradation of EOPs in recent years and, in order to reduce the risk of EOPs in water environment, non-thermal plasma has been a widely studied and applied technology. Summaries of some representative studies are compiled in Table 1.

properties: persistence, lipophilicity, and long-distance mobility. These properties enable them to perform bioamplification and bioaccumulation in animals and seriously harm the


**Table 1.** Overview of work done in the degradation of EOPs in water by non-thermal plasma.

*Water* **2022**, *14*, 1351


**Table 1.** *Cont.*

Initial concentration: 20 ppm

*Water* **2022**, *14*, 1351


**Table 1.** *Cont.*


**Table 1.** *Cont.*

Non-thermal plasma technology has been proved to be an effective method to degrade EOPs in water environment. For example, Yang et al. [59] studied the degradation of bisphenol A (BPA) in water by dielectric barrier discharge (DBD) plasma. The results showed that BPA was completely degraded (100%) within 25 min when the discharge voltage reached 16.8 kV. Satisfactory BPA degradation performance was achieved in a relatively short treatment time, which proved the superiority of non-thermal plasma treatment. Aggelopoulos et al. [91] investigated the degradation of enrofloxacin (ENRO) in aqueous solution in a gas-liquid nanosecond-pulsed dielectric barrier discharge (NSP-DBD) plasma reactor. Under the optimal pulse voltage and pulse frequency, ENRO was degraded completely (100%) after 20 min, and the corresponding energy yield was 1.1 g/kWh.

3,30 ,4,40 -tetrachlorobiphenyl (PCB77) was selected as the target pollutant for DBD treatment in order to verify the effectiveness of non-thermal plasma in the degradation of POPs in aqueous solution [89]. Their study showed that non-thermal plasma can effectively degrade PCB77 in aqueous solution. After DBD plasma treatment for 2 min, more than 75% of PCB77 was degraded. In addition, the biotoxicity of PCB77 degradation products was also evaluated, and it was found that DBD degradation products of PCB77 are almost non-toxic, which demonstrated that non-thermal plasma is an efficient, green, and environmentally friendly treatment technology to remove POPs from the environment.

The discharge voltage applied in non-thermal plasma significantly affects the degradation/mineralization efficiency of EOPs. On the one hand, with the increase of discharge voltage, the intensity of ultraviolet radiation increases, which leads to the improvement of pollutant oxidation. On the other hand, the number of reactive species, especially •OH, increases with the discharge voltage, which may lead to enhanced degradation of pollutants. However, as the discharge voltage increases excessively, the energy yield generally decreases. It is possible that the increased discharge voltage leads to increased energy waste, indicating that more electrical energy is being converted to heat [92,93]. Therefore, reasonable control of discharge voltage is the key of non-thermal plasma technology.

The pH value of aqueous solution is also one of the key factors for the degradation of EOPs by non-thermal plasma technology. It not only affects the properties of EOPs, but also affects the generation of reactive species in the non-thermal plasma system, which is probably the main reason why the results of different studies seem contradictory. The pH dependence of degradation efficiency has been extensively studied by many researchers. In general, the formation of •OH is more intense under neutral or alkaline conditions. Some previous reports also support that some reactive species (e.g., H2O<sup>2</sup> and O3) can decompose more quickly under alkaline conditions, forming •OH, leading to higher efficiency of pollutant degradation [94–96]. However, in some studies, it was found that better degradation efficiency was achieved under acidic conditions. In the study of Li et al. [97], as the pH value of the solution increased from 2.0 to 10.0, the reaction rate constants of tetracycline (TC), sulfadiazine (SD) and ciprofloxacin (CIP) decreased by 40.9%, 60.0%, and 65.0%, respectively. The lower degradation efficiency under alkaline conditions can be explained by the deprotonation of pollutant molecules, which may lead to the contraction of the bond length of pollutant molecules, thus increasing the stability of molecular structure and enhancement of hydrophilicity of pollutant molecules, thus reducing the interaction with reactive species. In addition, in the relatively high pH environment, the generated •OH will be quenched by OH– [98], thus inhibiting the degradation of pollutants. Therefore, pH is not a simple parameter, and the optimal pH value for different studies may be different, and largely depends on the chemical structure and properties of pollutants.

The reaction of EOPs with different molecular structures to oxidative attacks may be different and largely depends on the substituents on the benzene ring. Generally speaking, EOPs with stable molecular structure have strong resistance to reactive species oxidation. Li et al. [97] investigated the degradation of three antibiotics with different substituents and chemical properties by non-thermal discharge plasma oxidation, namely TC, SD, and CIP. The results showed that the three antibiotics could degrade effectively, but the reaction kinetics were different. The authors speculated that the significant difference in degradation performance of the three antibiotics may be due to their different molecular structures. To verify this experimental conclusion, the authors further determined the relationship between the chemical structure of these antibiotics and their removal efficiency by using Gaussian calculations. The ionization potential (IP) of organic compounds was calculated using the following equation (Equation (1)). Compared with SD and CIP, TC had the lowest ionization potential and was therefore more easily oxidized by reactive oxygen species (ROS). Therefore, organic pollutants with different structures have different degradation effects, which was also found in the study of Kim et al. [31]. Unfortunately, they did not explain the specific reasons. Banaschik et al. [77] explained why some pharmaceutical compounds are more recalcitrant than others. Aromatic ring systems, unsaturated bonds,

and electron donating functional groups (+I/+M) increased molecular reactivity towards plasma treatment and also towards other AOPs that are relying on the generation of •OH.

$$\text{IP} = 1.3124 \times (-\varepsilon \text{HOMO}) + 0.514 \text{ eV} \tag{1}$$

### *2.2. Soil Remediation*

At present, there is more and more research on remediation of EOPs contaminated soil by non-thermal plasma technology, and the excellent treatment effect has gradually attracted the attention of researchers. Summaries of some representative studies are presented in Table 2.

Non-thermal plasma has been proved to be effective in repairing EOPs-contaminated soil. Aggelopoulos et al. [101] studied the degradation of atrazine (ATZ) in soil by DBD discharge plasma, and the results showed that ATZ with initial concentration of 100 and 10 mg/kg could be degraded in dry soil with a degradation efficiency of 86.9% and 98.1% after plasma treatment for 60 min, respectively. Lou et al. [105] studied the remediation of chloramphenicol (CAP)-contaminated soil by DBD plasma. The results showed that the degradation efficiency of CAP was close to 81% after 20 min of plasma treatment, which demonstrated the feasibility of non-thermal plasma in removing pharmaceutical compounds from soil. Li et al. [112] used pulsed DBD plasma system to rehabilitate phenanthrene (PHE)-contaminated soil. Under the condition of 0.6 L/min air flow and 110 V voltage, the removal efficiency can reach 87.3% within 20 min, and the energy yield is 0.01 mg/kJ.

The researchers found that water content in soil is one of the most important factors affecting the removal efficiency of non-thermal plasma. A certain amount of water molecules can promote the production of •OH [114]. Wang et al. [100] demonstrated that in dry soil (0% moisture), about 65.6% of glyphosate was degraded after 45 min of DBD plasma treatment, and when soil moisture increased to 10%, the proportion increased to 86.5% over the same treatment time. However, soil moisture increased further to 20%, while glyphosate degradation efficiency decreased to 76.5%. The researchers attributed this to the fact that as the water content increased, the pores of the soil became clogged with water molecules, resulting in reduced transport of reactive species through the soil, resulting in less effective degradation of pollutants. In addition, some researchers have suggested that the presence of water contributes to this by helping dissolve organic matter in soil particles, allowing the dissolved organic matter to compete with the target pollutant for reactive species [115].

Different types of power supplies used to drive various plasma reactors also have a great impact on the efficiency of soil remediation. Recently, more and more researchers have focused on the plasma driven by nanosecond pulse power supply. Nanosecond pulsed plasmas have the following advantages: (1) higher electron energy [116,117], since most of the energy in the discharge process is used to accelerate the electron energy rather than neutral gas; (2) stronger chemical activity [118], because the nanosecond pulse discharge can produce more high-energy electrons, the inelastic collision in the discharge process is more intense, and more reactive species can be produced, the chemical activity is stronger than other discharges such as alternating current (AC) and direct current (DC); (3) better uniformity and stability [119,120]. Aggelopoulos et al. [106] studied the remediation of CIPcontaminated soil by nanosecond-pulsed DBD plasma. Under the optimal conditions (pulse voltage 17.4 kV, pulse frequency 200 Hz), CIP was completely degraded in soil (~99%), and treatment time was only 3 min, with a corresponding energy efficiency of 4.6 mg/kJ, which is quite high for soil remediation. However, the plasma discharge in this study was carried out in the gas phase above the soil surface, which would lead to poor permeability of ultraviolet radiation and reactive species produced by plasma in the soil, thus affecting the treatment effect. Hatzisymeon et al. [107] designed a new type of discharge plasma reactor to alleviate this problem. In such reactors, the reactive species produced by the discharge and plasma were produced directly in the contaminated medium, rather than in the gas

phase above the contaminated medium. Under optimized conditions, the energy efficiency of the system was 21.2 mg/kJ. In addition, this system could degrade CIP almost completely over a wide range of soil thicknesses (2.4 to 9.4 mm). At the actual concentration of CIP contamination in soil (20 mg/kg), the degradation process was very rapid and complete. At high initial contaminant concentrations (200 mg/kg), considerable degradation efficiency was also achieved. Water content of up to 10% did not appear to significantly affect process efficiency, which is important for implementation under practical conditions.


**Table 2.** Overview of work done in the degradation of EOPs in soil by non-thermal plasma.



*2.3. Comparison with Other AOPs*

Comparisons between non-thermal plasma and other AOPs are difficult because experimental conditions (such as molecular structure, initial concentration, treatment

volume, etc.) vary greatly and these parameters significantly affect the degradation process. In order to find a more accurate evaluation of degradation efficiency and energy yield, Hama Aziz et al. [74,121] focused on the degradation of 2,4-dichlorophenoxyacetic acid (2,4-D), 2,4-dichlorophenol (2,4-DCP), diclofenac (DCF), and ibuprofen (IBP) by several AOPs: ozonation, photocatalysis, and non-thermal plasma. The common reactor design of all experiments can directly compare the degradation efficiency and energy yield obtained by different methods. The specific comparison results are shown in Table 3. Comparing these AOPs from the perspective of degradation efficiency and energy yield, it is found that there is an obvious gap in the degradation of pollutants with different molecular structures, which indicates that different AOPs will be affected by the molecular structure.


**Table 3.** Comparison of EOPs degradation by non-thermal plasma and other AOPs.

Compared with photocatalysis, ozonation, and non-thermal plasma have the advantages of fast degradation rate, high degradation efficiency, and high energy yield. It can also be seen from Table 3 that ozonation and non-thermal plasma show similar performance. Taking 2,4-D as an example, the energy yield of non-thermal plasma is 8.8 g/kWh, while the energy yield of ozonation for the removal of 2,4-D is 6.6 g/kWh. It is worth noting that the mineralization efficiency of pollutants by ozonation is relatively low, and very good mineralization can be obtained by non-thermal plasma. Li et al. [122] compared the degradation of IBP by different AOPs. Except that the energy yield of photo-Fenton is slightly higher than that of water film DBD plasma, the energy yield of other AOPs is relatively low. However, photo-Fenton takes a longer time and the removal efficiency is not very high.

### **3. Mechanism of Emerging Organic Pollutants Degradation by Non-Thermal Plasma** *3.1. Reactive Species in Non-Thermal Plasma Discharges*

The ability of non-thermal plasma to produce highly reactive species in situ is well known. Their formation is mainly triggered by the collision of high-energy electrons produced in the discharge with gas atoms or molecules. Once the discharge process has occurred, reactive species can also be produced by radical recombination reactions or de-excitation of metastable substances [96]. The most abundant primary and secondary species formed in liquid or gas-liquid environments are hydroxyl radical (•OH), ozone (O3), and hydrogen peroxide (H2O2), which are associated with the degradation of target pollutants. However, many other ROS and RNS are produced in plasma, and can also contribute to the decomposition of pollutants, such as singlet oxygen (1O2), atomic oxygen (O), superoxide anion radical (•O<sup>2</sup> – ), peroxide hydroxyl radical (HO2•), nitrite (NO<sup>2</sup> – ), nitrate (NO<sup>3</sup> – ), peroxynitrite (ONOO– ), etc. The redox potentials of common oxidants are shown in Figure 2. These reactive species react with pollutants in water or soil, or high-energy electrons in an electric field react directly with pollutants, degrading them into small intermediates and further splitting them into carbon dioxide and water molecules. In

addition to oxidizing species, reductive species in discharge plasma may also contribute to the degradation of pollutants in water, such as aqueous electron (*E* <sup>0</sup> <sup>=</sup> <sup>−</sup>2.77 V) and H• radicals (*E* <sup>0</sup> <sup>=</sup> <sup>−</sup>2.30 V). Furthermore, some physical effects, such as ultraviolet radiation, heat, and shock wave, are often accompanied in the plasma discharge process. In conclusion, the degradation of organic compounds by discharge plasma depends mainly on the reactive species, while other physical effects are beneficial to the degradation process. H• radicals (*E*0 = −2.30 V). Furthermore, some physical effects, such as ultraviolet radiation, heat, and shock wave, are often accompanied in the plasma discharge process. In conclusion, the degradation of organic compounds by discharge plasma depends mainly on the reactive species, while other physical effects are beneficial to the degradation process.

small intermediates and further splitting them into carbon dioxide and water molecules. In addition to oxidizing species, reductive species in discharge plasma may also contribute to the degradation of pollutants in water, such as aqueous electron (*E*0 = −2.77 V) and

*Water* **2022**, *14*, x FOR PEER REVIEW 13 of 29

**Figure 2.** Comparison of oxidation potential of reactive species by non-thermal plasma. **Figure 2.** Comparison of oxidation potential of reactive species by non-thermal plasma.

•OH is the second most reactive substance after fluorine atom, and they attack most organic pollutant molecules with a rate constant of 106–109 M−1 s−1, which is 106–1012 times faster than ozone [123,124]. The production of •OH in water or plasma in contact with water was discussed in detail [125]. Bruggeman et al. found that the production of •OH depends largely on plasma parameters, such as gas temperature (Tg), electron temperature (Te), ionization degree, electron and ion density, and gas composition. The results showed that •OH formed by electron dissociation (Equation (2)) and dissociation attachment (Equation (3)) of water molecules are dominant in plasma with Te higher than 2 eV. However, when Te is between 1 and 2 eV, electron-ion dissociation recombination (Equa-•OH is the second most reactive substance after fluorine atom, and they attack most organic pollutant molecules with a rate constant of 106–10<sup>9</sup> M−<sup>1</sup> s −1 , which is 106–10<sup>12</sup> times faster than ozone [123,124]. The production of •OH in water or plasma in contact with water was discussed in detail [125]. Bruggeman et al. found that the production of •OH depends largely on plasma parameters, such as gas temperature (Tg), electron temperature (Te), ionization degree, electron and ion density, and gas composition. The results showed that •OH formed by electron dissociation (Equation (2)) and dissociation attachment (Equation (3)) of water molecules are dominant in plasma with Te higher than 2 eV. However, when Te is between 1 and 2 eV, electron-ion dissociation recombination (Equations (4) and (5)) and ion-ion dissociation recombination (Equations (6) and (7)) also play an important role in the mass production of •OH. Even at high enough ionization degree, it is also the main formation pathway of •OH.

$$\text{e}^- + \text{H}\_2\text{O} \rightarrow \text{HO} \\ \text{o} + \text{H} + \text{e}^- \tag{2}$$

$$\text{e}^- + \text{H}\_2\text{O} \rightarrow \text{HO} \\ \text{\bullet} + \text{H}^- \tag{3}$$

$$\text{e}^- + \text{H}\_2\text{O}^+ \rightarrow \text{HO}\bullet + \text{H}\bullet \tag{4}$$

$$\text{e}^- + \text{H}\_3\text{O}^+ \rightarrow \text{HO} \\ \text{o} + \text{H}\_2 + \text{e}^- \tag{5}$$

$$\rm H^{-} + H\_{2}O^{+} \rightarrow HO\bullet + H\_{2} \tag{6}$$

In the case of air or oxygen as the feed gas, O3 is formed by non-thermal plasma, and

the process has been well documented [126]. The electrons generated in the discharge excite and dissociate diatomic oxygen (Equation (8)), and the resulting atomic oxygen reacts

eି + HଷOା → HO • +Hଶ + eି (5)

Hି + HଷOା → HO • + Hଶ +H• (7)

Hି + HଶOା → HO • + Hଶ (6)

$$\text{H}^+ + \text{H}\_3\text{O}^+ \rightarrow \text{HO}\bullet + \text{H}\_2 + \text{H}\bullet \tag{7}$$

In the case of air or oxygen as the feed gas, O<sup>3</sup> is formed by non-thermal plasma, and the process has been well documented [126]. The electrons generated in the discharge excite and dissociate diatomic oxygen (Equation (8)), and the resulting atomic oxygen reacts with another oxygen molecule in the presence of a third object (M: M is a third collision partner: O2, O3, O) to form ozone (Equation (9)).

$$\text{O}\_2 + \text{e}^- \rightarrow 2\text{O} \bullet + \text{e}^- \tag{8}$$

$$\rm O\bullet + O\_2 + M \rightarrow O\_3 + M \tag{9}$$

It is generally believed that H2O<sup>2</sup> is mainly formed by the dimerization of •OH in discharge in contact with water (Equation (10)):

$$\bullet \text{OH} + \bullet \text{OH} \to \text{H}\_2\text{O}\_2 \tag{10}$$

H2O<sup>2</sup> is a relatively stable oxidant and can accumulate in the liquid phase during plasma discharge. Locke and Shih [127] summarized the literature on the formation of H2O<sup>2</sup> in water using various discharge techniques and experimental conditions. The results showed that the efficiency of H2O<sup>2</sup> production depends largely on the experimental device, and the maximum energy yield is 80 g/kWh. In addition to the direct generation of •OH in the plasma, additional •OH may be generated by the interaction of dissolved O<sup>3</sup> with H2O<sup>2</sup> (Equation (11)).

$$\rm H\_2O\_2 + O\_3 \rightarrow O\_2 + HO\bullet + HO\_2\bullet \tag{11}$$

Formation of more •OH is particularly beneficial for degradation because •OH is a powerful non-selective oxidant that reacts with most organic compounds, including short-chain carboxylic acids, and complex intermediates produced during degradation of organic molecules. Thus, increased concentrations of •OH can ensure further degradation/mineralization of organic pollutants.

In the study of plasma discharge technology using air/N<sup>2</sup> as raw gas, in addition to ROS, RNS are also formed, such as nitric oxide (NO), nitrogen dioxide (NO2), nitrite (NO<sup>2</sup> <sup>−</sup>), nitrate (NO<sup>3</sup> <sup>−</sup>), and peroxynitrite (ONOO−) [95]. Even though N<sup>2</sup> is a very stable molecule with high bonding energy, high-energy electrons generated during plasma discharge can dissociate the molecule from atomic nitrogen (Equation (12)). Nitrogen oxides are then rapidly formed by the interaction of atomic nitrogen with diatomic or triatomic oxygen (Equations (13) and (14)). The exchange between the two nitrogen oxides can be achieved by the interaction of nitric oxide with ozone and the photodissociation of nitrogen dioxide by ultraviolet radiation produced in the plasma (Equations (15) and (16)). Further dissolution and oxidation of nitrogen oxides in aqueous media lead to the formation of NO<sup>2</sup> −, NO<sup>3</sup> −, and ONOO−.

$$\text{N}\_2 + \text{e}^- \rightarrow 2\text{N}\bullet + \text{e}^- \tag{12}$$

$$\rm N\bullet + O\_2 \rightarrow \rm NO\bullet + O\bullet \tag{13}$$

$$\rm{N}\bullet + \rm{O}\_{3} \rightarrow \rm{NO}\bullet + \rm{O}\_{2} \tag{14}$$

$$\text{NO} \bullet + \text{O}\_3 \rightarrow \text{NO}\_2 + \text{O}\_2 \tag{15}$$

$$\rm NO\_2 + hv \rightarrow \rm NO\bullet + \rm O\bullet \tag{16}$$

### *3.2. Degradation Pathways of Emerging Organic Pollutants*

Different degradation efficiencies and rates have been recorded using non-thermal plasma techniques for different classes of EOPs. This phenomenon is mainly due to the different chemical structure of EOPs. Different kinds of organic compounds require different amounts of energy to break bonds. Therefore, the complexity of degradation pathway is also a key factor in the final degradation results. During non-thermal plasma treatment, in addition to the residue of the target pollutant, some intermediates also exist in the water/soil, which will compete with the target pollutant for reactive species, resulting in poor degradation. However, little is known about the differences in degradation mechanisms and kinetics. Therefore, the degradation pathways of some different types of EOPs are summarized below. in poor degradation. However, little is known about the differences in degradation mechanisms and kinetics. Therefore, the degradation pathways of some different types of EOPs are summarized below.

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### 3.2.1. Selected EDCs: Bisphenol A

Based on most relevant research, the results suggested that •OH, O<sup>3</sup> and •NO<sup>2</sup> play an important role in the degradation of BPA and most intermediates. Generally, there are two main pathways for the degradation of BPA in non-thermal plasma [56–59], as shown in Figure 3. On the one hand, •OH reacts with phenolic hydroxyl groups in BPA to form TP1 through hydrogen abstraction. The C atom on the ortho position of the hydroxyl group of TP1 undergoes hydroxylation and a recombination reaction under the attack of •OH and •NO<sup>2</sup> to generate TP2 and TP3. Then, TP4 and TP5 are generated by tautomerization of the keto and enol forms. On the other hand, the C atom between the two benzene rings is cleaved under the oxidation of •OH and O<sup>3</sup> to generate TP6 and TP7. In addition, the above intermediate products can be further oxidized and ring-opened to generate a series of small molecular organic compounds, and finally mineralized into CO<sup>2</sup> and H2O. 3.2.1. Selected EDCs: Bisphenol A Based on most relevant research, the results suggested that •OH, O3 and •NO2 play an important role in the degradation of BPA and most intermediates. Generally, there are two main pathways for the degradation of BPA in non-thermal plasma [56–59], as shown in Figure 3. On the one hand, •OH reacts with phenolic hydroxyl groups in BPA to form TP1 through hydrogen abstraction. The C atom on the ortho position of the hydroxyl group of TP1 undergoes hydroxylation and a recombination reaction under the attack of •OH and •NO2 to generate TP2 and TP3. Then, TP4 and TP5 are generated by tautomerization of the keto and enol forms. On the other hand, the C atom between the two benzene rings is cleaved under the oxidation of •OH and O3 to generate TP6 and TP7. In addition, the above intermediate products can be further oxidized and ring-opened to generate a series of small molecular organic compounds, and finally mineralized into CO2 and H2O.

**Figure 3.** The proposed BPA degradation pathway. **Figure 3.** The proposed BPA degradation pathway.

#### 3.2.2. Selected PPCPs: Ibuprofen 3.2.2. Selected PPCPs: Ibuprofen

Based on density functional theory (DFT), the spatial configuration of organic molecules and electron cloud density distribution can be obtained by molecular orbital calculation, which is helpful to predict the degradation behavior of organic pollutants. Generally, there are three common pathways for IBP to undergo degradation by non-thermal plasma, as shown in Figure 4. First, •OH can cause IBP to lose its carboxyl structure to generate the product TP1. TP1 is further deprotonated to form an intermediate product TP2 under the oxidation of •OH, O3 and other reactive species. Then, TP3 and TP4 are generated through hydroxylation and demethylation reactions. Secondly, the 10C in molecular structure of IBP is prone to hydroxylation to form TP5. Next, TP5 is deprotonated to form the product TP6, followed by demethylation and decarboxylation under the action of reactive species to generate TP7 and TP8 in turn. Li et al. [122] proposed the possible degradation pathway of IBP in DBD plasma based on DFT analysis. DFT was used to describe the molecular properties of IBP, which can reveal the potential of IBP in specific degradation reactions, as well as the optimal location of electrophilic or nucleophilic reactions in the molecule. The results showed that the 2 and 5 C positions of IBP are vulnerable Based on density functional theory (DFT), the spatial configuration of organic molecules and electron cloud density distribution can be obtained by molecular orbital calculation, which is helpful to predict the degradation behavior of organic pollutants. Generally, there are three common pathways for IBP to undergo degradation by non-thermal plasma, as shown in Figure 4. First, •OH can cause IBP to lose its carboxyl structure to generate the product TP1. TP1 is further deprotonated to form an intermediate product TP2 under the oxidation of •OH, O<sup>3</sup> and other reactive species. Then, TP3 and TP4 are generated through hydroxylation and demethylation reactions. Secondly, the 10C in molecular structure of IBP is prone to hydroxylation to form TP5. Next, TP5 is deprotonated to form the product TP6, followed by demethylation and decarboxylation under the action of reactive species to generate TP7 and TP8 in turn. Li et al. [122] proposed the possible degradation pathway of IBP in DBD plasma based on DFT analysis. DFT was used to describe the molecular properties of IBP, which can reveal the potential of IBP in specific degradation reactions, as well as the optimal location of electrophilic or nucleophilic reactions in the molecule. The results showed that the 2 and 5 C positions of IBP are vulnerable to attacks by electrophiles and nucleophiles. Therefore, the third degradation pathway of IBP begins with the attack of the 2 or 5 C atom by the •OH. TP9 and P10 are generated by the substitution of •OH,

and subsequent hydroxylation, demethylation, and deprotonation reactions may also occur to generate a series of intermediate products. Finally, the benzene ring undergoes a ringopening process to form formic, acetic, and oxalic acids. These organic acid molecules continued to mineralize into CO<sup>2</sup> and H2O. the substitution of •OH, and subsequent hydroxylation, demethylation, and deprotonation reactions may also occur to generate a series of intermediate products. Finally, the benzene ring undergoes a ring-opening process to form formic, acetic, and oxalic acids. These organic acid molecules continued to mineralize into CO2 and H2O.

to attacks by electrophiles and nucleophiles. Therefore, the third degradation pathway of IBP begins with the attack of the 2 or 5 C atom by the •OH. TP9 and P10 are generated by

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**Figure 4.** The proposed IBP degradation pathway. **Figure 4.** The proposed IBP degradation pathway.

#### 3.2.3. Selected POPs: PCB77 3.2.3. Selected POPs: PCB77

The dechlorination of PCBs has always been the most important step in their degradation process. Huang et al. [89] applied DBD non-thermal plasma for the degradation of PCB77, and its degradation process and common reaction sites are shown in Figure 5. Firstly, •OH causes the dechlorination reaction of PCB77 and generates a series of the dechlorination product TP1. Due to the different reactivity of different sites on the benzene ring, the selectivity of the dechlorination reaction usually depends on the position of the chlorine atom, following the order of para > meta > ortho [128]. Secondly, •OH can lead to a benzene cycloaddition reaction, forming multiple C-OH when all C-Cl break. Since the O atom in the phenolic hydroxyl group can undergo p-π conjugation with the benzene ring, the p electron cloud is transferred to the benzene ring, which increases the electron cloud density on the benzene ring and is easy to attack by electrophiles [129,130]. Numerous sites on the benzene ring are attacked by various reactive species in the plasma, and then partially cleaved to form the intermediate product TP2. The aromatic ring structure in TP2 may be attacked to form unstable products through electrophilic addition. Finally, the above intermediate products are degraded into smaller organic molecules, CO2 or inorganic salts. The dechlorination of PCBs has always been the most important step in their degradation process. Huang et al. [89] applied DBD non-thermal plasma for the degradation of PCB77, and its degradation process and common reaction sites are shown in Figure 5. Firstly, •OH causes the dechlorination reaction of PCB77 and generates a series of the dechlorination product TP1. Due to the different reactivity of different sites on the benzene ring, the selectivity of the dechlorination reaction usually depends on the position of the chlorine atom, following the order of para > meta > ortho [128]. Secondly, •OH can lead to a benzene cycloaddition reaction, forming multiple C-OH when all C-Cl break. Since the O atom in the phenolic hydroxyl group can undergo p-π conjugation with the benzene ring, the p electron cloud is transferred to the benzene ring, which increases the electron cloud density on the benzene ring and is easy to attack by electrophiles [129,130]. Numerous sites on the benzene ring are attacked by various reactive species in the plasma, and then partially cleaved to form the intermediate product TP2. The aromatic ring structure in TP2 may be attacked to form unstable products through electrophilic addition. Finally, the above intermediate products are degraded into smaller organic molecules, CO<sup>2</sup> or inorganic salts. *Water* **2022**, *14*, x FOR PEER REVIEW 17 of 29

**4. Non-Thermal Plasma and Other Technologies Cooperate to Degrade Emerging Or-**

research fields to make up for the shortcomings of a single system by using it in combination with other methods [131,132]. Non-thermal plasma is a kind of AOPs which integrates many factors, such as reactive species, high-energy electrons, ultraviolet radiation, and so on, which shows good characteristics in the process of dealing with EOPs. However, non-thermal plasma also has some disadvantages, such as low utilization efficiency of active components and ultraviolet radiation, low energy utilization efficiency, and so on [72,133]. In view of these limiting factors, the synergistic plasma technology of oxidants and catalysts will become one of development trends in the future. Another barrier to plasma is the low mass transfer efficiency of the resulting reactive species from the plasma to the phase where pollutants are present. Therefore, when designing a plasma reactor, it is necessary to consider not only the generation of these reactive species, but also their efficient transfer to the target pollutants. Recently, microbubbles (MBs) have been considered as an effective method to improve the efficiency of plasma reactors. However, only few studies have been carried out, and its mechanism is still unclear. Therefore, it is necessary to study the effects of MBs on improving gas-liquid mass transfer in water treatment and the development of other means to improve the mass transfer efficiency of reactive species from plasma to medium is also the focus of plasma research in the future.

tion because of its high oxidation activity and good adaptability to various EOPs. The

of •SO4− (30–40 µs) is much longer than that of •OH (10−3 µs), which is beneficial to improve the degradation efficiency of pollutants [134]. Ultraviolet radiation, local high temperatures, and hydrated electrons produced by non-thermal plasma can activate persulfate (PS) to form•SO4−, which helps improve the energy efficiency of the plasma system (Equations (17) and (18)) [135,136]. Tang et al. [65] studied the degradation of TC in water by gas surface discharge plasma-activated PS. With the increase of the PS dosage, the removal efficiency of TC in DBD plasma was also significantly improved. The calculated synergistic factor was 1.856, indicating that the addition of PS has an obvious synergistic effect. Wu et al. [133] found that the addition of peroxymonosulfate (PMS) into the DBD plasma can increase the degradation efficiency of benzotriazole (BTA) by 47%. In addition,

(2.5–3.1 V) is higher than that of •OH (1.9–2.7 V), and the half-life

) has attracted much atten-

In recent years, the AOPs based on sulfate radical (•SO4-

**Figure 5.** The proposed PCB77 degradation pathway. **Figure 5.** The proposed PCB77 degradation pathway.

*4.1. Oxidant* 

redox potential of •SO4-

### **4. Non-Thermal Plasma and Other Technologies Cooperate to Degrade Emerging Organic Pollutants in Environment**

Because most single methods have some defects, it has been fully proved in various research fields to make up for the shortcomings of a single system by using it in combination with other methods [131,132]. Non-thermal plasma is a kind of AOPs which integrates many factors, such as reactive species, high-energy electrons, ultraviolet radiation, and so on, which shows good characteristics in the process of dealing with EOPs. However, nonthermal plasma also has some disadvantages, such as low utilization efficiency of active components and ultraviolet radiation, low energy utilization efficiency, and so on [72,133]. In view of these limiting factors, the synergistic plasma technology of oxidants and catalysts will become one of development trends in the future. Another barrier to plasma is the low mass transfer efficiency of the resulting reactive species from the plasma to the phase where pollutants are present. Therefore, when designing a plasma reactor, it is necessary to consider not only the generation of these reactive species, but also their efficient transfer to the target pollutants. Recently, microbubbles (MBs) have been considered as an effective method to improve the efficiency of plasma reactors. However, only few studies have been carried out, and its mechanism is still unclear. Therefore, it is necessary to study the effects of MBs on improving gas-liquid mass transfer in water treatment and the development of other means to improve the mass transfer efficiency of reactive species from plasma to medium is also the focus of plasma research in the future.

### *4.1. Oxidant*

In recent years, the AOPs based on sulfate radical (•SO<sup>4</sup> −) has attracted much attention because of its high oxidation activity and good adaptability to various EOPs. The redox potential of •SO<sup>4</sup> <sup>−</sup> (2.5–3.1 V) is higher than that of •OH (1.9–2.7 V), and the half-life of •SO<sup>4</sup> <sup>−</sup> (30–40 <sup>µ</sup>s) is much longer than that of •OH (10−<sup>3</sup> <sup>µ</sup>s), which is beneficial to improve the degradation efficiency of pollutants [134]. Ultraviolet radiation, local high temperatures, and hydrated electrons produced by non-thermal plasma can activate persulfate (PS) to form•SO<sup>4</sup> −, which helps improve the energy efficiency of the plasma system (Equations (17) and (18)) [135,136]. Tang et al. [65] studied the degradation of TC in water by gas surface discharge plasma-activated PS. With the increase of the PS dosage, the removal efficiency of TC in DBD plasma was also significantly improved. The calculated synergistic factor was 1.856, indicating that the addition of PS has an obvious synergistic effect. Wu et al. [133] found that the addition of peroxymonosulfate (PMS) into the DBD plasma can increase the degradation efficiency of benzotriazole (BTA) by 47%. In addition, energy production increased by 84%. The improvement of BTA degradation efficiency may be attributed to the activation of PMS by plasma, which increases the formation of reactive species and •SO<sup>4</sup> −. In addition, some studies have shown that the introduction of PMS can realize the secondary utilization of O<sup>3</sup> (Equations (19)–(23)) [135].

$$\bullet \text{HSO}\_5^{-} \stackrel{\text{plasma}}{\rightarrow} \bullet \text{SO}\_4^{-} + \bullet \text{OH} \tag{17}$$

$$\text{S}\_2\text{O}\_8^{2-} \stackrel{\text{plasma}}{\rightarrow} \text{2} \bullet \text{SO}\_4^- \tag{18}$$

$$\rm{^0SO\_5^{2-}} + \rm{O\_3} \rightarrow \rm{^-O\_3SO\_5^-} \tag{19}$$

$$\rm{\text{--}}\rm{\text{O}\_3\text{SO}\_5^-} \rightarrow \rm{\text{\textbullet}\rm{SO}\_5^-} + \rm{\text{\textbullet}\rm{O}\_3^-} \tag{20}$$

$$\bullet \mathrm{SO}\_5^- + \mathrm{O}\_3 \to \bullet \mathrm{SO}\_4^- + \mathrm{O}\_2 \tag{21}$$

$$\bullet \mathrm{O}\_3^- \to \bullet \mathrm{O}^- + \mathrm{O}\_2 \tag{22}$$

$$\bullet \bullet^{-} + \text{H}\_{2}\bullet \rightarrow \bullet \text{OH} + \bullet \text{H}^{-} \tag{23}$$

In addition, oxidants, such as percarbonate (SPC), ferrate, and H2O2, were also used in non-thermal plasma. Tang et al. [137] explored the synergistic effect of SPC and plasma, and the results showed that the addition of SPC was beneficial to the production of H2O<sup>2</sup>

and the decomposition of O3. When SPC was 52.0 mol/L and voltage was 4.8 kV, the removal efficiency of TC could reach 94.3% at 20 mg/L. Xu et al. [138] found that low concentration (0.1–1.0 mmol/L) of H2O<sup>2</sup> could promote the degradation of norfloxacin (NOR) by plasma, while high concentration (1.0–2.0 mmol/L) of H2O<sup>2</sup> could inhibit the degradation of NOR. Sang et al. [139] compared the effect of PMS, SPC, and ferrate on the degradation of Orange G (OG) by DBD plasma, and found that the degradation efficiency of OG by ferrate plasma was the best. The possible mechanism of degradation of EOPs by non-thermal plasma combined with different oxidants is shown in Figure 6. *Water* **2022**, *14*, x FOR PEER REVIEW 19 of 29

**Figure 6.** Possible degradation mechanisms of EOPs combined with different oxidants: (**a**) PS, (**b**) SPC, (**c**) Fe (VI). **Figure 6.** Possible degradation mechanisms of EOPs combined with different oxidants: (**a**) PS, (**b**) SPC, (**c**) Fe (VI).

#### *4.2. Catalyst 4.2. Catalyst*

plasma.

#### 4.2.1. Homogeneous Catalyst 4.2.1. Homogeneous Catalyst

Fe2+ is the most commonly used homogeneous catalyst in combination with non-thermal plasma. The addition of Fe2+ can form a Fenton system with the non-thermal plasma and produce more •OH to degrade EOPs (Equations (24) and (25)) [140,141]. Hao et al. [142] introduced iron ion (Fe2+/Fe3+) into the pulsed discharge plasma system. The experimental results showed that the addition of iron ion (Fe2+/Fe3+) could greatly improve the removal efficiency of 4-chlorophenol. At the same time, they found that the promoting effect of Fe2+ was greater than that of Fe3+. Xu et al. [138] found that a low concentration of Fe2+ can promote the degradation of NOR by DBD plasma, while a high concentration can inhibit the degradation of NOR, which may be due to the reaction of excessive Fe2+ with •OH to form Fe3+ and OH- (Equation (26)). At the same time, the Fe3+ can also react with H2O2 in aqueous solution (Equation (25)), thus reducing the concentration of H2O2. HଶOଶ + Feଶା → • OH + OHି + Feଷା (24) HଶOଶ + Feଷା → Hା + HOଶ • +Feଶା (25) Fe2+ is the most commonly used homogeneous catalyst in combination with nonthermal plasma. The addition of Fe2+ can form a Fenton system with the non-thermal plasma and produce more •OH to degrade EOPs (Equations (24) and (25)) [140,141]. Hao et al. [142] introduced iron ion (Fe2+/Fe3+) into the pulsed discharge plasma system. The experimental results showed that the addition of iron ion (Fe2+/Fe3+) could greatly improve the removal efficiency of 4-chlorophenol. At the same time, they found that the promoting effect of Fe2+ was greater than that of Fe3+. Xu et al. [138] found that a low concentration of Fe2+ can promote the degradation of NOR by DBD plasma, while a high concentration can inhibit the degradation of NOR, which may be due to the reaction of excessive Fe2+ with •OH to form Fe3+ and OH- (Equation (26)). At the same time, the Fe3+ can also react with H2O<sup>2</sup> in aqueous solution (Equation (25)), thus reducing the concentration of H2O2.

$$\mathrm{H\_2O\_2 + Fe^{2+} \to \cdot OH + OH^- + Fe^{3+}}$$

$$\mathrm{H\_2O\_2 + Fe^{3+} \to H^+ + HO\_2\bullet + Fe^{2+}}\tag{25}$$

$$\mathrm{Fe^{2+} + \bullet OH \to OH^{-} + Fe^{3+}}$$

In addition, the oxidizability of O<sup>3</sup> produced by non-thermal plasma is lower than that of •OH, while the introduction of Fe2+ can react with O<sup>3</sup> to form •OH (Equations (27) and (28)), thus improving the •OH content and energy efficiency of the non-thermal plasma. Feଶା + Oଷ → Oଶ + FeOଶା (27)

$$\text{Fe}^{2+} + \text{O}\_3 \rightarrow \text{O}\_2 + \text{FeO}^{2+} \tag{27}$$

$$\rm H\_2O + FeO^{2+} \rightarrow Fe^{3+} + \bullet OH + OH^- \tag{28}$$

#### 4.2.2. Heterogeneous Catalyst Non-thermal plasma co-heterogeneous catalytic oxidation technology refers to the

Non-thermal plasma co-heterogeneous catalytic oxidation technology refers to the addition of solid catalyst (generally used in powder form) in the discharge region of the plasma. Catalysts react with ultraviolet radiation or reactive species such as ozone and hydrogen peroxide generated in plasma discharge process, and then trigger a series of chain reactions to promote the generation of active free radicals and degradation of organic matter. Furthermore, the combined adsorption and catalytic action of catalysts with a large specific surface area can contribute to the removal of pollutants. Compared with other advanced oxidation technologies, plasma discharge has a variety of physical and chemical effects, such as high electric field, shock wave, etc., which can clean the surface of accelerators and contribute to the regeneration of catalysts [143]. At present, the solid catalysts used for plasma catalytic oxidation mainly include carbon catalysts represented by activated carbon, photocatalysts represented by TiO2, other metal catalysts, metal or metal oxide catalysts supported on supports, etc. The possible mechanism of degradation of EOPs by non-thermal plasma combined with heterogeneous catalyst is shown in Figure 7. addition of solid catalyst (generally used in powder form) in the discharge region of the plasma. Catalysts react with ultraviolet radiation or reactive species such as ozone and hydrogen peroxide generated in plasma discharge process, and then trigger a series of chain reactions to promote the generation of active free radicals and degradation of organic matter. Furthermore, the combined adsorption and catalytic action of catalysts with a large specific surface area can contribute to the removal of pollutants. Compared with other advanced oxidation technologies, plasma discharge has a variety of physical and chemical effects, such as high electric field, shock wave, etc., which can clean the surface of accelerators and contribute to the regeneration of catalysts [143]. At present, the solid catalysts used for plasma catalytic oxidation mainly include carbon catalysts represented by activated carbon, photocatalysts represented by TiO2, other metal catalysts, metal or metal oxide catalysts supported on supports, etc. The possible mechanism of degradation of EOPs by non-thermal plasma combined with heterogeneous catalyst is shown in Figure 7.

**Figure 7.** Possible degradation mechanisms of EOPs combined with heterogeneous catalyst. **Figure 7.** Possible degradation mechanisms of EOPs combined with heterogeneous catalyst.

Different from the homogeneous catalyst, the heterogeneous catalyst is easy to recover and separate; so, there are many studies on the coupling of the heterogeneous catalyst and plasma to degrade EOPs. Guo et al. [69] coupled pulsed discharge plasma (PDP) with Fe3O4 to promote the degradation of CAP. In PDP system, Fe3O4 not only catalyzed H2O2 to form a Fenton reaction, but also catalyzed O3, which promotes the formation of •OH. When the addition amount of Fe3O4 is 0.26 g/L, under the conditions of higher peak voltage, lower initial solution concentration, and lower initial pH value, CAP is beneficial to decomposition and has the best catalytic performance. Cheng et al. [144] used α-MnO2, β-MnO2, and γ-MnO2 to degrade CIP wastewater by DBD plasma-catalytic combined process. The results showed that the combination of DBD plasma and α-MnO2 has the highest degradation efficiency of CIP, and the degradation efficiency could reach 93.1%, which was 10.8% and 18.1% higher than that of β-MnO2 and γ-MnO2 catalyst in the plasma cat-Different from the homogeneous catalyst, the heterogeneous catalyst is easy to recover and separate; so, there are many studies on the coupling of the heterogeneous catalyst and plasma to degrade EOPs. Guo et al. [69] coupled pulsed discharge plasma (PDP) with Fe3O<sup>4</sup> to promote the degradation of CAP. In PDP system, Fe3O<sup>4</sup> not only catalyzed H2O<sup>2</sup> to form a Fenton reaction, but also catalyzed O3, which promotes the formation of •OH. When the addition amount of Fe3O<sup>4</sup> is 0.26 g/L, under the conditions of higher peak voltage, lower initial solution concentration, and lower initial pH value, CAP is beneficial to decomposition and has the best catalytic performance. Cheng et al. [144] used α-MnO2, β-MnO2, and γ-MnO<sup>2</sup> to degrade CIP wastewater by DBD plasma-catalytic combined process. The results showed that the combination of DBD plasma and α-MnO<sup>2</sup> has the highest degradation efficiency of CIP, and the degradation efficiency could reach 93.1%, which was 10.8% and 18.1% higher than that of β-MnO<sup>2</sup> and γ-MnO<sup>2</sup> catalyst in the plasma

alytic system, respectively. The photocatalyst TiO2 can enhance the degradation of EOPs

catalytic system, respectively. The photocatalyst TiO<sup>2</sup> can enhance the degradation of EOPs in the plasma system by utilizing the ultraviolet radiation produced by the plasma (Equations (29) and (30)) [145]. Lee et al. [83] found that the addition of TiO<sup>2</sup> increased the decomposition rate of DMP, but the excess dosage of TiO<sup>2</sup> resulted in ultraviolet radiation blocking and decreased the decomposition rate of DMP. Jogi et al. [146] found that the addition of TiO<sup>2</sup> can increase the amount of O<sup>3</sup> produced by the DBD plasma, which may be one of the reasons for improving the efficiency of pollutant degradation.

$$\text{TiO}\_2 + \text{h}\nu \rightarrow \text{TiO}\_2(\text{e}\_{\text{cb}}^- + \text{h}\_{\text{vb}}^+) \tag{29}$$

$$\rm H\_2O + h^+\_{vb} \to \rm \bullet OH + H^+ \tag{30}$$

A single metal oxide photocatalyst has a high recombination rate for electron-hole pairs generated by light energy and is only sensitive to ultraviolet light, while catalysts such as Fe3O<sup>4</sup> can use H2O<sup>2</sup> and O<sup>3</sup> but cannot use ultraviolet light [64,71]. Therefore, it is an ideal choice to prepare composite catalysts from metal oxides and other catalysts [147]. In order to make full use of the ultraviolet light generated by the plasma, Fe3O<sup>4</sup> was supported on reduced graphene oxide (rGO) [71]. Compared with using Fe3O<sup>4</sup> alone, rGO-Fe3O<sup>4</sup> further improved the degradation efficiency and kinetic constant of ofloxacin in the discharge plasma system. After 60 min of treatment, the degradation efficiency and kinetic constant reached 99.9% and 0.108 min−<sup>1</sup> , respectively.

In addition, composite catalysts composed of metal oxides and metal oxides, metal oxides, and metal elements are also used in non-thermal plasma. Ansari et al. [68] combined a ZnO/α-Fe2O<sup>3</sup> composite catalyst with DBD plasma to degrade antibiotic amoxicillin (AMX). The results showed that the ZnO/α-Fe2O<sup>3</sup> composite catalyst increased the degradation efficiency of AMX from 75.0% (sole DBD plasma) to 99.3% under optimal conditions. Wang et al. [148] found that compared with the plasma process, the combination of plasma and Mn/γ-Al2O<sup>3</sup> catalyst significantly improved the degradation efficiency of tetracycline hydrochloride. Under the discharge power of 1.3 W, the degradation efficiency of tetracycline hydrochloride could reach 99.3%, while the degradation efficiency of the plasma treatment was only 69.7%.

At present, plasma-catalytic systems are mainly focused on water remediation, but there are few reports about soil remediation. However, it has been demonstrated that plasma-catalytic systems improve the efficiency of plasma systems in soil remediation. Wang et al. [99,149] investigated the degradation of p-nitrophenol (PNP) in soil using a pulsed discharge plasma-TiO<sup>2</sup> catalytic system. Compared with the single plasma system, the system showed higher degradation performance of PNP. Increased TiO<sup>2</sup> content promoted PNP degradation to a certain extent, while further increased TiO<sup>2</sup> content had negative effects. At higher TiO<sup>2</sup> content, particles aggregation may reduce the interface area between contaminants and catalyst surface sites, thereby reducing the number of active sites on the catalyst surface and resulting in reduced PNP degradation [150].

In conclusion, the combination of plasma discharge and catalyst reveals the catalytic promoting effect on pollutant removal and improves the efficiency of non-thermal plasma.

### *4.3. Microbubbles*

MBs generally refer to bubbles with equivalent diameters less than 50 µm. Compared with millimeter-sized bubbles, MBs have some special properties that can enhance the discharge effect. Firstly, the gas-liquid mass transfer ability of MBs is strong. Unlike millimeter-sized bubbles, MBs will not rise rapidly from the water to the liquid surface and break. The rising speed of MBs in water is slower than that of millimeter-sized bubbles, and the residence time in water is longer, which greatly prolongs the gas-liquid contact time [151,152]. Figure 8 shows the characteristics of millimeter-sized bubbles and micronsized bubbles in water. Secondly, MBs can stimulate the generation of free radicals in the process of fragmentation. Adiabatic compression occurs during bubble contraction, which generates local high temperature and high pressure around the bubble. Such conditions

stimulate the decomposition of water molecules around the bubble into free radicals [153]. Takahashi et al. [152] demonstrated free-radical generation from the collapse of MBs in the absence of a harsh dynamic stimulus. Electron spin-resonance spectroscopy confirmed freeradical generation by the collapsing MBs. The increase of the surface charges (ζ potentials) of the MBs, which were measured during their collapse, supported the hypothesis that the significant increase in ion concentration around the shrinking gas-water interface provided the mechanism for radical generation. stimulate the decomposition of water molecules around the bubble into free radicals [153]. Takahashi et al. [152] demonstrated free-radical generation from the collapse of MBs in the absence of a harsh dynamic stimulus. Electron spin-resonance spectroscopy confirmed free-radical generation by the collapsing MBs. The increase of the surface charges (ζ potentials) of the MBs, which were measured during their collapse, supported the hypothesis that the significant increase in ion concentration around the shrinking gas-water interface provided the mechanism for radical generation.

**Figure 8.** Comparison of millimeter-sized bubbles and micron-sized bubbles in water. **Figure 8.** Comparison of millimeter-sized bubbles and micron-sized bubbles in water.

Studies showed that compared with plasma treatment alone, the presence of MBs can significantly improve the treatment efficiency, and preliminary studies speculate that the presence of MBs can improve gas-liquid mass transfer in plasma water treatment [154,155]. MBs were introduced with different carrier gases (air, N2 and Ar) in the needleplate pulsed discharge reactor to enhance the interface reaction [156]. Due to the surface ζ-potential, MBs can effectively enrich pollutants, and the large specific surface area also leads to a high gas-liquid interface area, which enhances the plasma reaction with pollutants. At the same time, these unique physical properties can also promote the mass transfer from gas to liquid in the system. Wang et al. [157] evaluated the degradation of ATZ in aqueous solution by DBD/MBs/PS system in order to develop a more efficient and environmentally friendly PS activation method. The observed ATZ removal efficiency (DBD/MBs/PS > DBD/PS > MBs/PS) confirmed the synergistic effect of DBD/MBs/PS. Based on the ATZ removal efficiencies of 64% and 56% in DBD/MBs/PS and DBD/PS systems, the mass transfer contribution rate was calculated as 13% in DBD/MBs/PS system. The synergies of DBD/MB/PMS systems are largely due to the interaction between DBD, MBs, and PMS. DBD plasma produces a large number of reactive species when gas molecules dissociate and form high-energy electrons. By combining DBD plasma and MBs, the plasma can be ignited inside the MBs, where an electron avalanche breakdown occurs and continues until the MBs crashes. Studies showed that compared with plasma treatment alone, the presence of MBs can significantly improve the treatment efficiency, and preliminary studies speculate that the presence of MBs can improve gas-liquid mass transfer in plasma water treatment [154,155]. MBs were introduced with different carrier gases (air, N<sup>2</sup> and Ar) in the needle-plate pulsed discharge reactor to enhance the interface reaction [156]. Due to the surface ζ-potential, MBs can effectively enrich pollutants, and the large specific surface area also leads to a high gas-liquid interface area, which enhances the plasma reaction with pollutants. At the same time, these unique physical properties can also promote the mass transfer from gas to liquid in the system. Wang et al. [157] evaluated the degradation of ATZ in aqueous solution by DBD/MBs/PS system in order to develop a more efficient and environmentally friendly PS activation method. The observed ATZ removal efficiency (DBD/MBs/PS > DBD/PS > MBs/PS) confirmed the synergistic effect of DBD/MBs/PS. Based on the ATZ removal efficiencies of 64% and 56% in DBD/MBs/PS and DBD/PS systems, the mass transfer contribution rate was calculated as 13% in DBD/MBs/PS system. The synergies of DBD/MB/PMS systems are largely due to the interaction between DBD, MBs, and PMS. DBD plasma produces a large number of reactive species when gas molecules dissociate and form high-energy electrons. By combining DBD plasma and MBs, the plasma can be ignited inside the MBs, where an electron avalanche breakdown occurs and continues until the MBs crashes.

### **5. Practical Implications of This Study**

**5. Practical Implications of This Study**  Among a large number of research works in the past decade, non-thermal plasma technology has been proved to be a promising environmental remediation technology. Among a large number of research works in the past decade, non-thermal plasma technology has been proved to be a promising environmental remediation technology. However, there are still some challenges to be solved and breakthroughs to be made.

However, there are still some challenges to be solved and breakthroughs to be made. (1) One of the most serious challenges of non-thermal plasma technology is how to improve the energy yield of the treatment system and reduce the operation cost. According to the summary of existing research, the energy yields of different research results vary greatly, covering several orders of magnitude. In addition to the different molecular structures of pollutants, the design and experimental conditions of a nonthermal plasma reactor also have a great influence. Due to the particularity of in situ (1) One of the most serious challenges of non-thermal plasma technology is how to improve the energy yield of the treatment system and reduce the operation cost. According to the summary of existing research, the energy yields of different research results vary greatly, covering several orders of magnitude. In addition to the different molecular structures of pollutants, the design and experimental conditions of a nonthermal plasma reactor also have a great influence. Due to the particularity of in situ generation of reactive species by non-thermal plasma, the main improvement of

reactor design mainly lies in how to maximize the generation of reactive species and effectively transfer them to the medium where pollutants exist.


Some suggestions for future development direction are also put forward.


### **6. Conclusions**

This review provides readers with a comprehensive overview of non-thermal plasma in environmental remediation. Based on the above discussion of this review, some important conclusions can be drawn.

Firstly, the research progress of non-thermal plasma on EOPs (i.e., EDCs, PPCPs, and POPs) in different environmental media is summarized. It is found that many parameters will affect the degradation efficiency of non-thermal plasma, and it is clear from previous studies that non-thermal plasma technology shows the potential for effective degradation and complete mineralization of EOPs, which cannot be successfully removed by conventional treatment methods. For soil remediation, direct ignition of plasma discharge in the interior of soil pores seems to promote the penetration of UV and short-lived reactive species in the soil matrix, thus further promoting effective remediation of the entire porous medium.

Secondly, the generation of reactive species and the degradation mechanism of EOPs in non-thermal plasma system are discussed. Similar to other AOPs, non-thermal plasma also emphasizes that •OH plays a central role in the degradation of pollutants. The degradation pathways of various EOPs are mainly the following reactions: •OH is directly generated in discharge or further decomposed by O<sup>3</sup> and H2O2. Then, according to the molecular structure of organic pollutants (e.g., electron density), the oxidative degradation pathway may also include the reaction with O3.

Finally, some measures to improve the efficiency of non-thermal plasma are comprehensively summarized. In terms of water remediation, the formation of MBs under the water surface seems to have significant advantages, because it is a more effective means to provide reactive species to degrade EOPs in water and the existence of MBs enhances the

dissolution and mass transfer of reactive species generated by plasma from gas phase to liquid phase. For soil remediation, the combination of non-thermal plasma and catalyst seems to be a promising technology to break through the bottleneck of single plasma.

In conclusion, non-thermal plasma technology has a good application prospect in environmental remediation.

**Author Contributions:** All authors contributed extensively to the work presented in this paper. Y.H.: investigation, visualization, writing—original draft; W.S.: writing—review and editing; W.L.: writing—review and editing; W.Z. and C.Z.: resources; D.J.: formal analysis. All authors have read and agreed to the published version of the manuscript.

**Funding:** This work was supported by the National Natural Science Foundation of China project 51108360.

**Data Availability Statement:** Not applicable.

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

### **References**


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