*Article* **Simultaneous Phosphate Removal and Power Generation by the Aluminum–Air Fuel Cell for Energy Self-Sufficient Electrocoagulation**

**Xiaoyu Han 1,2, Hanlin Qi 1, Youpeng Qu 3, Yujie Feng 2,\* and Xin Zhao 1,\***


#### **Featured Application: The Al–air fuel cell provides an energy self-sufficient electrocoagulation system for phosphate removal.**

**Abstract:** A self-powered electrocoagulation system with a single-chamber aluminum–air fuel cell was employed for phosphate removal in this study. Electricity production and aluminum hydroxides in solution were also investigated. When the NaCl concentration increased from 2 mmol/L to 10 mmol/L, the phosphate removal increased from 86.9% to 97.8% in 60 min. An electrolyte composed of 10 mmol/L of NaCl was shown to obtain a maximum power density generation of 265.7 mW/m2. When the initial solution pH ranged from 5.0 to 9.0, 98.5% phosphate removal and a maximum power density of 338.1 mW/m<sup>2</sup> were obtained at pH 6.0. Phosphate was mainly removed by aluminum hydroxide adsorption. These results demonstrate that the aluminum–air fuel cell can be applied as electricity-producing electrocoagulation equipment. Aluminum–air fuel cells provide an alternative method to meet the goal of carbon neutrality in wastewater treatment compared with traditional energy-consuming electrocoagulation systems.

**Keywords:** aluminum–air fuel cell; electrocoagulation; phosphate removal; electricity production

#### **1. Introduction**

The discharge of phosphorus into water bodies has caused a serious worldwide water eutrophication problem [1,2]. A concentration of 0.1–0.2 mg/L phosphate can induce incipient eutrophication in running water, while the critical concentration reduces to 0.005–0.01 mg/L of phosphate for still water [3]. The eutrophication of water threatens the health of aquatic creatures, livestock, and even humans. Therefore, the removal of phosphorus from municipal wastewater is essential to avoid water eutrophication. In China, the national limit for total phosphorus (TP) effluent of the Class 1A Discharge Standard is 0.5 mg/L [2]. The TP concentration is approximately 4–9 mg/L in China's municipal wastewater treatment plants [4]. Biological phosphorus removal processes are the conventional and most widely applied technology. However, the biological treatment system occupies a large area of land, requires high energy consumption, and needs to control the operation conditions to reach this standard [5]. For the high discharge standard of phosphate in municipal wastewater treatment, physicochemical treatments, such as chemical precipitation by adding aluminum or iron coagulants, are applied for the further removal of phosphate after biological treatment [6,7]. Chemical coagulation increases the operating costs, produces excessive amounts of chemical sludge, and leads to potential risks of secondary pollution.

**Citation:** Han, X.; Qi, H.; Qu, Y.; Feng, Y.; Zhao, X. Simultaneous Phosphate Removal and Power Generation by the Aluminum–Air Fuel Cell for Energy Self-Sufficient Electrocoagulation. *Appl. Sci.* **2023**, *13*, 4628. https://doi.org/10.3390/ app13074628

Academic Editor: Apostolos Giannis

Received: 8 February 2023 Revised: 27 March 2023 Accepted: 28 March 2023 Published: 6 April 2023

**Copyright:** © 2023 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/).

Electrocoagulation (EC) is considered a promising wastewater treatment technology that comprises three aspects: coagulation, flotation, and electrochemistry [8]. In the EC process, a sacrificial metal anode, such as Al or Fe, dissolves and releases metal ions into the solution with an impressed current [8]. Various metal hydroxides are formed by the hydrolysis process. Hydrogen bubbles and OH− ions are discharged at the cathode. Pollutants and impurities are mainly removed by multiple physicochemical reactions with metal hydroxides and generate precipitates simultaneously [9,10]. EC is an effective method for phosphate removal, and the advantages of the EC system compared with chemical coagulation are its easy operation, no additional chemical requirements, and less sludge production [10–12]. However, the electrical energy demands need to be reduced to make the EC process economically viable and eco-friendly. Recently, a new type of EC process using an air cathode has been investigated to reduce EC energy demands [13–15]. At the air cathode, an oxygen reduction reaction proceeds with a related catalyst, and oxygen gas is obtained through air diffusion without aeration. With a minor added operation voltage of 0.5 V, 98% of phosphate was removed in a shortened time of 15 min [16]. The reason for this decrease in the electrical energy requirement of the EC with the air cathode is that the metal anode and the oxygen reduction air cathode exist in a potential gradient.

Metal–air fuel cells, which consist of a metal anode and an air cathode, have been investigated for synchronous contaminant removal and energy generation [17]. Over the last few decades, metal–air fuel cells have been implemented to remove arsenate, chromium, p-arsanilic acid, and humic acid, and for algal recovery [18–24]. Some researchers have also applied metal–air fuel cells for phosphate removal and recovery. When wastewater contained nitrogen and phosphorus together, Mg–air fuel cells were employed to recycle the nitrogen and phosphorus by forming struvite precipitation [25,26]. Iron–air fuel cells are a new option for recovering phosphate from wastewater. Phosphate was recovered by the formation of vivianite precipitation. Within 3 months of continuous operation, the iron–air fuel cell output voltage achieved approximately 400 mV, and 97% of phosphate was removed from synthetic industrial wastewater [27]. Different configurations of iron–air fuel cells were constructed to treat wastewater containing anaerobically digested sludge. The phosphate removal rate of a two-chamber fuel cell obtained a rate of 11.60 mg-P/L/h [28]. Aluminum is also commonly used as an electrode in electrocoagulation for phosphate removal, and aluminum–air fuel cells have been utilized to remove arsenate. Thus, it is important to research the performance of phosphate removal and the influence of the phosphate removal rate on aluminum–air fuel cells.

In this study, an energy self-sufficient single-chamber aluminum–air fuel electrocoagulation system was constructed for phosphate removal. Different electrolyte concentrations, initial pH values of the solution, and phosphate concentrations were investigated to determine their influences on phosphate removal and power generation. The coagulated precipitation in an aluminum–air fuel cell was also selected and analyzed to demonstrate the phosphate removal process in the aluminum–air fuel cell. The self-powered aluminum–air fuel cell provides an energy-friendly electrocoagulation process for nutrient contaminant removal, which is significant in wastewater treatment for achieving the goal of carbon neutrality.

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

#### *2.1. Reactor Construction*

An aluminum–air fuel cell reactor was constructed using a rectangular plexiglass reactor of 80 mL (length: 4 cm, width: 2 cm, height: 10 cm, Figure 1). The area of the air cathode was 40 cm<sup>2</sup> (4 cm × 10 cm). The air cathode, including stainless steel mesh, activated carbon (Xinsen Carbon Co., Ltd., Nanping, Fujian, China), conductive carbon black (Jinqiushi Chemical Co. Ltd., Tianjin, China), and PTFE (60 wt%, Hesen Electrical Co., Ltd., Shanghai, China), was constructed using a rolling process, as described in a previous study [29]. For the diffusion layer, conductive carbon black and an appropriate amount of ethanol were mixed using mechanical and ultrasonic agitation for 15 min. The PTFE was

then added and the solution was further stirred for 15 min. The carbon black was kneaded like dough. Then, it was compacted and shaped between the rollers, and the diffusion layer was pressed to a thickness of 0.4 mm. Finally, the diffusion layer sheet and the stainless-steel mesh were rolled together and calcined in a muffle furnace at 340 ◦C for 20 min. Activated carbon powder was used as the catalyst to create the catalytic layer. Similar to the operation of the diffusion layer, a certain amount of activated carbon powder, ethanol, and PTFE were mixed and rolled to create a catalytic layer sheet with a thickness of 0.3 mm. The catalytic layer was then rolled together on the other side of the stainless-steel mesh. The anode was a piece of aluminum mesh of 40 cm2 (mesh size 1.0 mm × 1.5 mm), which was placed opposite to the cathode. Before operation, the aluminum mesh was washed with an acetone solution to remove the residual oil and stains on the anode surface. The anode was then immersed in 1.0 mol/L HCl solution for 10 min to remove the oxidation layer on the surface. The anode was used after it was washed with deionized water.

**Figure 1.** Schematic diagram (**a**) and photo (**b**) of the aluminum–air fuel cell for electrocoagulation.

#### *2.2. Solutions and Experimental Procedure*

The phosphate solution was dispensed with KH2PO4 (99.9%) and NaCl (99.5%), which were prepared to increase the solution's conductivity. The reactor was operated with a resistor of 10 Ω. The concentration of NaCl in the solution varied from 2 mmol/L to 10 mmol/L with 5 mg/L PO4 <sup>3</sup>−-P to analyze the solution conductivity effect. The initial solution pH of 5.0–9.0 was adjusted by adding 1.0 mol/L HCl or 1.0 mol/L NaOH. The PO4 <sup>3</sup>−-P concentrations ranged from 1 mg/L to 5 mg/L to compare the performances with different original PO4 <sup>3</sup>−-P concentrations. Approximately 1.5 mL of solution was removed every 5 min for PO4 <sup>3</sup>−-P analysis. The self-powered electrocoagulation experiments were conducted at room temperature.

#### *2.3. Measurements and Calculations*

The samples extracted at different times were tested after they were filtered through 0.45 μm pore diameter syringe filters. The phosphate concentration was measured using the ammonium molybdate spectrophotometric method (photoLab® 7600 UV, WTW, Munich, Germany) [30]. The pH was tested using a pH meter (pH 7110, WTW, Munich, Germany). The voltage across the resistor was monitored at 1 min intervals using a data acquisition system (PISO-813, ICPDAS Co., Ltd, Taiwan, China). The current density (I) was calculated based on Ohm's law (I = U/(RA)), and the power density (P) was calculated as P = IU/A, where U is the voltage (mV), R is the external resistor (Ω), and A is the surface area of the cathode. Polarization curves and electrode potential curves were obtained by recording the voltages under various external resistances (5–1000 Ω). A saturated calomel electrode (SCE, 0.242 V vs. a standard hydrogen electrode) was inserted into the reactor chamber as the reference electrode to test the electrode potential. The theoretical dissolved aluminum ion

concentration was calculated, and the number of aluminum ions was calculated according to Faraday's law. The equation for Faraday's law is as follows:

$$\mathbf{m} = \frac{\mathbf{I} \times \mathbf{A} \times \mathbf{t} \times \mathbf{M}}{\mathbf{Z} \times \mathbf{F}} \tag{1}$$

Here, m is the theoretical dissolved mass of aluminum ions (g), I is the current density (A/cm2), t is the operation time (s), A is the surface area of the cathode (40 cm2), M is the molecular weight of the aluminum material (M = 26.98 g/mol), Z is the number of electrons released from the aluminum anode (the Z of aluminum is 3 eq/mol), and F is Faraday's constant (96,485 C/mol). The volume of the electrolyte solution used for the aluminum ion concentration calculation was V = 80 cm3. The P-precipitants obtained from the 10 mmol/L NaCl solution were collected and vacuum-dried at 40 ◦C for chemical phase analysis. The precipitates were tested by X-ray diffraction (D8 ADVANCE, Bruker, Karlsruhe, Germany) under operating conditions of 100 mA and 40 kV in the range of 2θ = 10–90◦. The aluminum anode and precipitates were analyzed using scanning electron microscopy (Carl Zeiss AG, Sigma500, Oberkochen, Germany) and energy-dispersive spectroscopy.

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

#### *3.1. Effects of Electrolyte Concentration*

The electrolyte concentration of NaCl directly influenced the resistance between the electrodes and changed the output current [30]. In the EC process, increasing the electrolyte concentration could increase the phosphate removal efficiency and reduce the electrolysis operation time. The aluminum–air fuel cell system's phosphate removal performance was monitored with different concentrations of NaCl ranging from 10 mmol/L to 2 mmol/L with a resistor of 10 Ω. With the decrease in the electrolyte concentration, the phosphate removal rate in the aluminum–air fuel cell showed a downward trend, and the required reaction time was prolonged. The phosphate removal in the aluminum–air fuel cell containing 8 mmol/L of NaCl was slightly lower than that of the cell containing 10 mmol/L of NaCl. When the NaCl concentration decreased to 2 mmol/L, the phosphate removal reached 97.8%, 93.3%, 91.8%, and 86.9% at 60 min, respectively (Figure 2a). White flocculent precipitates were observed with different operating concentrations of electrolytes. At 10 mmol/L of NaCl, the phosphate removal was 98.0% at 50 min, and the rate of removal decreased with longer operation times, which demonstrated that the operation time could be reduced with a higher electrolyte concentration. When the NaCl concentration was decreased from 10 mmol/L to 2 mmol/L, the average output current density of the aluminum–air fuel cell decreased from 0.20 mA/cm2 to 0.06 mA/cm2 (Figure 2b). The low conductivity of the solution led to an increase in the internal resistance of the fuel cell, slowing down the mass transfer in solution and the dissolution rate of the aluminum anode [24]. When the NaCl concentration declined from 10 mmol/L to 2 mmol/L, the calculated dissolutions of aluminum ions were 33.7 mg/L, 32.2 mg/L, 20.3 mg/L, and 9.5 mg/L, respectively. A higher conductivity electrolyte increased the anode dissolution and produced larger amounts of aluminum hydroxides for complex precipitation reactions in the solution; thus, a higher electrolyte concentration was favorable for phosphate removal [20]. In a traditional EC system, an increase in the electrolyte concentration can improve energy consumption, leading to an increase in the EC operating costs. On the contrary, a higher electrolyte concentration will generate more electricity for the aluminum–air fuel cell.

**Figure 2.** Effects of different electrolyte concentrations on (**a**) phosphate removal and (**b**) current generation in the aluminum–air fuel cell.

To analyze the influence of electrolyte concentrations on the electrochemical performance of the aluminum–air fuel cell, polarization curves and electrode potentials were tested with different electrolyte concentrations. As the electrolyte conductivity decreased, the maximum power density of the aluminum–air fuel cell reduced from 265.7 mW/m2 to 76.6 mW/m<sup>2</sup> (Figure 3a). The electrode potentials of the aluminum anode and the air cathode both decreased with a decline in conductivity (Figure 3b). With a higher electrolyte concentration, more aluminum ions dissolved, generating a higher current output.

**Figure 3.** (**a**) Power density and polarization curves and (**b**) electrode potential curves for the aluminum−air fuel cell with different electrolyte concentrations.

#### *3.2. Effects of Initial pH*

The pH value can be a significant factor in the electrocoagulation process. In this work, the initial solution pH ranged from 5.0 to 9.0, the initial phosphate concentration was 5 mg/L, and the NaCl concentration was 10 mmol/L. As shown in Figure 4a, different initial pH values influenced the phosphate removal rate in the aluminum–air fuel cell system. When the solution pH was alkaline, the phosphate removal rate was faster than the rate under acidic conditions in the early stage. Within 50 min, all phosphate removal rates reached values higher than 90% with different initial pH values. At 60 min, as the initial pH 5.0 changed to pH 9.0, the phosphate removal reached 96.3%, 98.5%, 96.3%, 97.0%, and 95.6%. In contrast to the removal rate trend, the highest level of final phosphate removal was obtained at pH 6.0, which was likely due to the higher aluminum dissolution with a higher current output at pH 6.0 [31]. The average current density of the aluminum–air fuel cell ranged between 0.18 and 0.20 mA/cm2, which changed slightly during electricity production. At pH values of 5.0 to 9.0, the calculated aluminum ion concentrations were 30.7 mg/L, 33.9 mg/L, 33.1 mg/L, 31.9 mg/L, and 32.2 mg/L. These results also indicated that the self-driven electrocoagulation could adapt to different pH conditions, as shown in previous research [20].

**Figure 4.** Effects of different initial pH values on (**a**) phosphate removal and (**b**) current generation in the aluminum−air fuel cell.

The power density of the aluminum–air fuel cell varied slightly under different pH conditions (Figure 5a,b). The maximum power densities of the aluminum–air fuel cell were higher at pH 6.0 and pH 7.0, being 338.1 mW/m<sup>2</sup> and 332.8 mW/m2, respectively. Under alkaline conditions, the cell maximum power densities decreased; at pH 9.0, the maximum power density was 292.8 mW/m2. The difference in the fuel cell electrode potential under different pH conditions was not obvious.

**Figure 5.** (**a**) Power density and polarization curves and (**b**) electrode potential curves for the aluminum−air fuel cell with different initial pH.

The electrocoagulation performance was influenced by the aluminum dissolution and the final pH. In this experiment, pH variation was also investigated with different initial pH values, as shown in Table 1. After 60 min of operation with the initial pH value ranging from 5.0 to 9.0, the final pH changed in different trends. For the acidic and neutral conditions, the pH value increased, while with initial pH values of 8.0 and 9.0, the final pH value decreased. The variation trends were similar to previously reported trends [20]. The dissolution rate of aluminum was also related to the current density and the pH value. The behaviors of the electrode and aluminum ions in solution also affected the solution pH. For the aluminum–air fuel cell, the electrode and overall reactions were as shown in Equations (2)–(4). At the aluminum anode, electrons were released from the aluminum, and then the aluminum ions reacted with water, as shown in Equation (5) [32,33]:

$$\text{Anode}: \text{Al} \rightarrow \text{Al}^{3+} + \text{3e}^- \text{E}^0 = 1.66 \text{ V} \tag{2}$$

$$\text{Cathode:}\,\text{:}\,\text{O}\_2 + 4\text{H}^+ + 4\text{e}^- \to 2\text{H}\_2\text{O} \\ \text{E}^0 = 1.23\text{ V} \tag{3}$$

$$\text{Overall:}\\\text{Al} + \text{O}\_2 + \text{H}^+ \rightarrow \text{Al}^{3+} + \text{H}\_2\text{O}\\\text{E}^0 = 2.89 \text{ V} \tag{4}$$

$$\text{Al}^{3+} + 3\text{H}\_2\text{O} \rightarrow \text{Al(OH)}\_3 + 3\text{H}^+ \tag{5}$$


**Table 1.** The pH value changes of the aluminum–air fuel cell before and after operation for 60 min.

At the air cathode, OH− ions can also be generated through the four-electron oxygen reduction reaction with water, as shown in Equation (6):

$$\rm O\_2 + 2H\_2O + 4e^- \rightarrow 4OH^- \tag{6}$$

At pH 5–6, for most of the polymeric species, the ratios of precipitation of Al3+ and OH− were less than 3, which caused the OH− ions to accumulate in the solution during the operation process and increased the pH [33]. Under alkaline conditions, more OH− ions could be consumed by aluminum ions to form polymeric aluminum hydroxide, and this process decreased the pH value.

#### *3.3. Effects of Initial Phosphate Concentration*

Generally, the removal efficiency of phosphate was negatively correlated with the initial phosphate concentration in the EC process, and the phosphate removal was limited at the start of the operation, as the amount of coagulants produced was relatively small [14, 34,35]. In this study, the phosphate concentration was set to approach the concentration of real domestic wastewater. A 10 mmol/L NaCl electrolyte and a 10 Ω operating condition were used, and the initial concentrations of phosphate in the simulated wastewater were set to 5 mg/L, 4 mg/L, 3 mg/L, 2 mg/L, and 1 mg/L, respectively. As the results in Section 3.1 show, the operation time could be reduced to 50 min with 10 mmol/L of NaCl. The maximum phosphate removal rate was achieved at 50 min, and the level of phosphate removal decreased after 60 min. At lower phosphate concentrations, a higher phosphate removal rate was obtained at the beginning of the operation. With the increase in the phosphate concentration, the phosphate removal rate increased at 50 min, achieving 92.2%, 93.0%, 92.6%, 96.9%, and 98.0%, respectively (Figure 6). In the actual sewage treatment process, the concentration of phosphate in sewage changes greatly. These results indicate that in a relatively low phosphate concentration range, the aluminum–air fuel cell system can achieve a high phosphate removal efficiency. Compared with the existing EC process, which consumes electricity, an electrical-power-generating electrocoagulation system was applied to this work, which provided an energy-friendly and promising electrocoagulation process for phosphate removal.

**Figure 6.** The relationship between the initial phosphate concentration and phosphate removal in the aluminum−air fuel cell.

#### *3.4. Comparison of Energy Demand and Implications*

In traditional EC systems for phosphorous removal using aluminum electrodes or a combination of aluminum and iron electrodes, the energy demand per cubic meter of P-wastewater was approximately 1.1–6.1 kWh, as shown in Table 2. When an air cathode was applied as the electrode, the energy consumption decreased compared with the traditional EC system. In the air cathode EC system which used a titanium inert electrode, only 0.009–0.06 kwh was added, and the operation time required was only 15 min [13]. Photovoltaic solar modules were also applied to power the EC process, and the energy from the solar energy modules was enough for the EC system, although the solar energy conversion efficiency was less than 13% [30]. For the aluminum–air fuel cell, with a 10 mmol/L NaCl electrolyte and a 10 Ω operation condition, the average power density at 60 min was approximately 160 mW/m2. As the power density curves show, the aluminum–air fuel cell output power density was related to the output current. The power production was less than that of the solar cell; however, the aluminum–air fuel cell electrocoagulation system did not require an external electricity unit such as a solar cell module. This simple device is more practical in wastewater treatment, and solar energy is also highly dependent on weather. Compared with the traditional EC system, the operating cost for anode consumption is similar, but the cost of the air cathode in the aluminum–air fuel cell is higher than that of the conventional EC cathode [19]. The air cathode cost and lifetime should be one of the future research points considered for scale-up. Therefore, in future studies, the construction of a scale-up system, the electrode lifetime, and overall economic assessment including labor, sludge handling, maintenance, and depreciation costs will be key issues for practical application [18,36–38].

**Table 2.** Energy consumption of different electrocoagulation systems and the removal rates of phosphorus reported in different studies with Al electrodes.


Another advantage of the metal–air fuel cell is that the air cathode is capable of generating hydrogen peroxide with a suitable oxygen reduction catalyst, which has the potential to treat wastewater by combining coagulation and oxidization processes. For domestic wastewater treatment, the aluminum–air fuel cell can be used at the pre-treatment segment and the advanced treatment segment to remove phosphate. Organic contaminants such as antibiotics, resistant genes, and persistent organic pollutants which are harmful to human health can coexist with phosphate in wastewater. The multiple physicochemical processes in aluminum–air fuel cells provide promising methods to treat the complex effluents in the advanced treatment stage.

#### *3.5. Mechanisms of Phosphate Removal*

The aluminum anode was analyzed before and after fuel cell operation using SEM to clarify the phosphate removal process (Figures 7 and 8). The anode surface formed a corrosion pit after the operation (Figure 7a,b), providing evidence that the anode spontaneously dissolved aluminum ions into the solution. Previous studies have shown that pitting corrosion exists on metal surfaces when they are covered by a compact oxide film [40]. Furthermore, chloride ions in the electrolyte solution are the aggressive ions on the metal surface. Chloride ions can trigger the pitting corrosion reaction, and the passivation layer on the electrode surface will decompose. For the Al anode, a corrosion-protective film consisting of aluminum oxide can easily form on the surface, and the protective film is stable within a pH range of 4–8.5. However, chloride ions can react with the protective film

and expose the anode surface, which will increase the anode dissolution rate [34]. These processes are shown in Equations (7)–(9) [41]:

$$\rm Al\_2O\_3 + Cl^- + 6H^+ \rightarrow 2AlCl\_3 + 3H\_2O \tag{7}$$

$$\text{Al(OH)}\_{3} + \text{nCl}^{-} \rightarrow \text{Al(OH)}\_{3-\text{n}}\text{Cl}\_{\text{n}} + \text{nOH}^{-}, \text{n} = 1, 2, 3 \tag{8}$$

$$\text{AlCl}\_3 + \text{Cl}^- \rightarrow \text{AlCl}\_4^+ \tag{9}$$

**Figure 7.** SEM results of the Al electrode in this study, (**a**) the original electrode, and (**b**) the electrode after the operation.

**Figure 8.** SEM images of phosphate precipitation in aluminum–air batteries (**a**) and EDS results (**b**).

The morphology of the precipitates from 10 mmol/L of NaCl and 5 mg/L of phosphate was observed by SEM, and the elemental compositions were analyzed by EDS, as shown in Figure 8a,b. The precipitates had a blocky structure, and the EDS results showed that the precipitates were mainly composed of P, Al, and O elements, which indicated that the phosphate was removed with the aluminum compounds.

There are two main mechanisms that exist in the aluminum coagulation phosphate removal process, namely co-precipitation by monomeric aluminum and adsorption by aluminum hydroxide [42]. In the EC process with aluminum electrodes, both direct precipitation and adsorption by aluminum hydroxide were observed, and direct precipitation for phosphate removal was more efficient than hydroxide adsorption removal [43,44]. The phosphate removal behavior was related to the water conditions and the aluminum species. AlPO4 precipitate can be formed by Al3+ and phosphate, and the Al13 species can also induce formation of AlPO4 by complex adsorption [45]. For the coagulation process, pH influences the speciation of aluminum and accordingly affects the phosphorus removal mechanism. When the pH value was below 3.5, the dominant aluminum species was Al3+. As the pH value increased from 3.5 to 10.0, the dominant species were complex monomeric and polymeric aluminum hydroxides. Al(OH)4 − became the predominant species when the pH value was above 10.0 [46]. The precipitated products from different initial pH values

were examined by XRD to reveal the crystalline structure (Figure 9). Three diffraction peaks were observed in the flocs. The second peak was in the range of 23–33◦, and the third peak was at approximately 41◦, which was likely within the spectrum of Al(OH)3. The crystal plane of AlPO4 was also found in the second peak range [45–47]. However, the absence of sharp peaks in the XRD results indicated that the flocs had a poor crystalline structure, and that many of the precipitates were in an amorphous state, which was similar to a previous report on aluminum coagulants [47]. Combining the precipitates, XRD results, and the aluminum species constructions with different pH values, it can be concluded aluminum hydroxide adsorption was the main mechanism for removing phosphate. A schematic diagram of the phosphate removal process is summarized in Figure 10.

**Figure 9.** XRD patterns of floc precipitation in the aluminum–air fuel cell.

**Figure 10.** Schematic diagram of the phosphate removal process in the aluminum−air fuel cell.

#### **4. Conclusions**

A single-chamber aluminum–air fuel cell system was utilized to remove phosphate, and the effects of various factors were examined for phosphate removal. The phosphate removal mechanism was also discussed. An increase in the electrolyte concentration improved the level of phosphate removal and electricity production. With 10 mmol/L of NaCl, 97.8% phosphate removal was achieved, and the maximum power density reached 265.7 mW/m2. For different initial pH values, the optimal condition for electricity production was pH 6.0, and 98.5% phosphate removal was achieved in 60 min. The average current density obtained was 0.20 mA/cm2, and the maximum power density was 338.1 mW/m2. Aluminum hydroxides were synthesized, and phosphate was primarily removed via the aluminum hydroxide adsorption process. The aluminum fuel cell can treat phosphate wastewater and generate energy simultaneously. Compared with conventional

electrocoagulation systems, the aluminum–air fuel cell system can be considered an energy self-sufficient alternative, which is significant for achieving carbon neutrality in the wastewater treatment industry.

**Author Contributions:** X.H.: data curation and writing—original draft preparation; H.Q.: writing original draft preparation; Y.Q.: methodology and supervision; Y.F.: conceptualization; X.Z.: writing review and editing. All authors have read and agreed to the published version of the manuscript.

**Funding:** This research was funded by the National Natural Science Foundation of China (No. 52204184) and the Fundamental Research Funds for the Central Universities (No. N2201019).

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

**Informed Consent Statement:** Not applicable.

**Data Availability Statement:** The data presented in this study are available on request from the corresponding authors.

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

#### **References**


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**Ran Jiang 1, Jianyu Fu 1, Zhaoyang Wang <sup>2</sup> and Cunku Dong 1,\***


**Abstract:** The electrocatalytic hydrogen evolution reaction (HER) of a given metal catalyst is intrinsically related to its electronic structure, which is difficult to alter for further improvement. Recently, it was discovered that the density of grain boundaries (GBs) is mechanistically of great importance for catalytic activity, implying that GBs are quantitatively correlated with the active sites in the HER. Here, by modeling the atomistic structure of GBs on a Au(110) surface, we find that HER performance is greatly enhanced by Au GBs, suggesting the feasibility of the HER mediated by GBs. The promoted HER performance is due to an increase in the capability of binding adsorbed hydrogen on the sites around GBs. A Au catalyst with a dominantly exposed (110) plane is synthesized, where considerable GBs exist for experimental verification. It is found that HER activity is inherently correlated with the density of the GBs in Au NPs. The improvement in HER activity can be elucidated from the geometrical and electronic points of view; the broken local spatial symmetry near a GB causes a decrease in the coordination numbers of the surface sites and the shift up of the d–band center, thereby reducing the limiting potential for each proton−electron transfer step. Our finding represents a promising means to further improve the HER activity of a catalyst.

**Keywords:** grain boundaries; electrocatalytic hydrogen evolution; density functional calculations; gold

#### **1. Introduction**

In recent decades, human beings have relied on fossil fuels for over 80% of their total energy needs. Hydrogen, as a clean, economic and renewable energy carrier, is considered to be an attractive alternative to traditional fossil fuels, and it can greatly alleviate the global greenhouse effect and energy crisis at present [1]. Amongst various hydrogen production techniques, the electrocatalytic hydrogen evolution reaction (HER) via electrochemical water splitting, an important energy recovery technique, has received great attention as a hot research topic [2–4]. Consequently, ever–increasing efforts have been devoted to developing a variety of new catalysts with the aim of improving HER performance [5,6]. However, noble metal–based materials such as platinum (Pt) (e.g., Pt/C) are still the most efficient catalysts widely used to catalyze hydrogen evolution, owing to their outstanding thermodynamical and kinetic features for the HER [3]. Unfortunately, Pt is an extremely scarce and precious metal, which results in a high cost for the HER. Therefore, designing a new catalyst or boosting the activity of existing non–noble materials other than Pt has become urgent.

As an important descriptor, hydrogen bonding ability is often taken into account in the screening or designing of HER catalysts [7–9]. According to the Sabatier principle, either a too weak or too strong adsorption ability results in a drop in HER performance in that a too weak adsorption of H\* leads to a pronounced decline in intermediate stabilization, while a too strong adsorption inhibits the desorption of gaseous hydrogen [10]. The superior HER of Pt is due to its optimal H\* binding energy, or Gibbs free energy, which is close to zero. To

**Citation:** Jiang, R.; Fu, J.; Wang, Z.; Dong, C. Grain Boundary—A Route to Enhance Electrocatalytic Activity for Hydrogen Evolution Reaction. *Appl. Sci.* **2022**, *12*, 4290. https:// doi.org/10.3390/app12094290

Academic Editor: Leonarda Francesca Liotta

Received: 30 March 2022 Accepted: 19 April 2022 Published: 24 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/).

date, many non–noble materials suffer from a weak H\* binding ability. Identifying how to enhance their H\* binding activity remains an enormous challenge. Therefore, various routes have been proposed to control the microstructure of catalysts in order to expose more active facets or sites [11–13]. However, further improving HER activity is still a great challenge due to the inherent electronic structure of the catalysts themselves, leading to a certain capacity for H\* binding, which almost cannot be tuned. Inspired by a previous study on the relationship between grain boundaries (GBs) and catalytic activity in various reactions [14–16], GBs rich in the atomic arrangement disorder were found to alter the local electronic structure around GBs [17,18]. In addition, GBs feature under–coordinated sites and microstrains [16,19], which favor H binding. Kim [17], using extensive density functional theory (DFT) calculations to model the atomistic structure of GBs on a Au(111) surface, concluded that the grain boundary of Au is conducive to the adsorption of COOH by the active site, thus enhancing electrochemical CO2 reduction. Dong [18] believes that the GB sites on a Au(110) surface lead to a high selectivity toward CH3OH. Thus, GBs are expected to have a stronger H\* binding capability so as to compensate for the unsaturated coordination of GB atoms resulting from such a disorder. However, these studies lack an investigation of the atomic and electronic structures of grain boundaries and an explanation of the enhanced catalytic performance of grain boundaries.

Herein, we first predict the impact of grain boundaries on a Au surface on the HER occurring at GBs via density functional theory (DFT) calculations, in light of the much weaker H\* binding capacity of Au [20]. Then, a Au NP catalyst with GBs is prepared for the verification of GB–assisted HER activity. HER activity greatly increases with an increase in the density of GBs in Au NPs. Additionally, the atomic and electronic structures are analyzed to determine the role of the GBs. On the basis of the theoretical and experimental studies, it is found that the GBs can prodigiously promote the chemical bonding ability of H\* intermediate, resulting in enhanced catalytic activity in hydrogen evolution.

#### **2. Compactional and Experimental Methods**

#### *2.1. Computational Model and Method*

Of many possible high–angle grain boundaries (HAGBs), an ∑6{2–21} HAGB atomistic model on a Au(110) surface was constructed using coincidence site lattice (CSL) theory, as shown in Figure 1. This HAGB surface model consists of three layers containing 52 Au atoms (Figure S1). Five atop catalytic active sites were selected, labeled as s1–s5 (Figure 1).

**Figure 1.** Top view of the Σ6{2–21} HAGB model on Au(110) surface with three atomic layers in an orthogonal supercell (25.29 <sup>×</sup> 8.94 <sup>×</sup> 17.80 Å3). The GB area is highlighted by blue lines. The selected Au around GB is labeled as s1–s5.

Spin–polarized DFT calculations were performed using Vienna ab initio Simulation Package (VASP) [21]. The revised Perdew−Burke−Ernzerhof (rPBE) exchange–correlation functional within the generalized gradient approximation (GGA) was used with the projector augmented wave (PAW) pseudopotential [22]. A plane–wave cutoff energy of 450 eV was used for clean surface and adsorbate surface relaxation. The relaxation was complete when the residual force was less than 0.05 eV/Å. During the structure relaxation, two downmost layers were fixed to their original position of the slab without the adsorbate, while the uppermost layer and adsorbate were fully relaxed. Then, 2 × 1 × 1 Monkhorst−Pack mesh sampling was employed for the clean surface and adsorbate surface in the surface

Brillouin zone. A vacuum space of 15 Å perpendicular to the slab surface was used to avoid artificial interaction. Additional computational details about the calculation of the free energy diagram can be found in the Supporting Information.

#### *2.2. Preparation of Au/CFP Electrodes*

Au nanoparticles were deposited on carbon fiber paper by magnetron sputtering under a sputtering vacuum of 5 × <sup>10</sup>−<sup>2</sup> mbar and a sputtering current of 20 mA for 30 s using a tabletop DC magnetron sputtering coater (Leica EM SCD 500). The purity of the Au target used for deposition was 99.999%. The annealing of Au/CFP electrodes was performed using a tube furnace (Thermo Scientific) with a flowing Ar atmosphere at 100 sccm, heated at a given temperature (200~400 ◦C) for 2 h. The use of an Ar atmosphere prevented the formation of carbon coatings on the Au nanoparticles.

#### *2.3. Electrochemical Characterization*

All electrochemical measurements were performed on a CHI 1100C electrochemical workstation (Chenhua, China). The Pb underpotential deposition (upd) measurements of the Au/CFP electrodes were performed in 0.1 M NaOH solution containing 1 mM Pb(OAc)2, and Pt and Ag/AgCl electrode (3.0 M KCl) were the counter electrode and reference electrode, respectively. The voltammogram scan rate was 50 mV s–1.

Electrochemical hydrogen evolution measurements were performed in 0.5 M H2SO4 with continuous purging of N2 (>99.999% purity) using a standard three–electrode cell, where the Au/CFP electrodes, the graphite rod (>99.999% purity) and the Hg/Hg2Cl2 electrode (saturated KCl solution) were used as the working, counter and reference electrodes, respectively. The potential was calibrated with respect to RHE in the high–purity hydrogen–saturated electrolyte, with a Pt plate as the working electrode. A flow of N2 was maintained over the electrolyte (0.5 M H2SO4) during the electrochemical measurements in order to eliminate the possible effects of other gases. The polarization curves were recorded with a scan rate of 5 mV s−<sup>1</sup> without *iR* correction.

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

#### *3.1. Theoretical Prediction of GB–Assisted HER*

#### 3.1.1. Gibbs Free Energy

The HER is a multi–step electrochemical process occurring on the surface of an electrode [23]. In acid media, two possible routes have been elucidated for the HER, i.e., the Volmer−Heyrovsky and the Volmer−Tafel mechanisms, which can be described as follows [24,25]:

$$\text{H}\_{3}\text{O}^{+} + \text{M} + \text{e}^{-} \rightleftharpoons \text{H} \ast - \text{M} + \text{H}\_{2}\text{O} \tag{1}$$

$$\text{NH} \; \text{\*}-\text{M} + \text{H}\_{3}\text{O}^{+} + \text{e}^{-} \rightleftharpoons \text{M} + \text{H}\_{2} + \text{H}\_{2}\text{O} \tag{2}$$

2H ∗ −M -2M + H2 (3)

Step (1) refers to the Volmer reaction, while steps (2) and (3) refer to the Heyrovsky and Tafel reactions, respectively. We first explore the impact of the GBs on the Au(110) surface on the Volmer reaction, which dominantly determines the overall HER performance; for this, we chose various active sites at and near the Au GBs (Figure 1). Figure 2a shows the Gibbs free energy change (ΔGH\*) profile of the HER in acid media. The ΔGH\* values of the flat Au(110) and Pt(111) GB–free surfaces are also presented for comparison (Tables S1 and S2). It should be noted that ΔGH\* dramatically reduces once the GBs are introduced. Additionally, ΔGH\* exhibits site–dependent HER activity for sites in the vicinity of the GBs. To be specific, the ΔGH\* for s1 is 0.18 eV, approximately 0.4 eV less than the flat Au(110) (0.58 eV), whereas s2 and s3 have a ΔGH\* of about −0.07 eV; this is less than that of Pt (−0.09 eV), which is more close to zero. As is well known, the optimal value of ΔGH\* for the HER is zero, as hydrogen binding is neither too strong nor too weak. Therefore, the activity for the Volmer reaction is greatly promoted by the GBs. In addition, the sites near the GBs (s4 and s5) reduce ΔGH\* by ~0.13 eV with regard to the Au(110) surface, suggesting

that GBs on the Au(110) surface not only greatly promote HER activity on their own (s1–s3) but also activate the sites nearby (s4–s5) to further lower the reaction barrier.

**Figure 2.** (**a**) The calculated free−energy diagram of HER at the equilibrium potential (U = 0 V) for Au(110) (black), Pt(111) (green) and different active sites on Au GBs. (**b**) The volcano relation of the limiting potential for HER as a function of EB(H\*). The arrow indicates the desired direction for catalyst design with higher activity, and the color pentagrams represent GB sites on Au surface for prodigious promotion of HER. Theoretical overpotentials (*η*HER) are the vertical difference between the points and the equilibrium potential (red dashed line).

#### 3.1.2. Limiting Potentials

To further evaluate the HER performance, we compared the limiting potential (UL = −ΔGH\*/e) of various metal materials with different exposed facets (Tables S3 and S4). Figure 2b presents a volcano plot, which typically correlates EB(H\*) with UL. GB–free Au(110) is located far from the top on the right side of the volcano plot, indicating that Au(110) binds H\* weakly enough not to form a stable intermediate, which readily desorbs off the surface; thus, a much higher overpotential is required to yield gaseous hydrogen [10], which is in agreement with the previous discussion. When GBs are formed on Au(110), HER activity moves close to the top of the volcano. The active sites at the GBs (s1–3) possess superior HER performance to almost all the metal catalysts considered here. Additionally, the UL required is also reduced for sites near the GBs (s4 and s5). In particular, s1, s2 and s3 lie close to the top of the volcano plot near Pt, although PDS is different, suggesting that their HER activity is similar to that of the Pt extended surface. The enhanced HER performance is mainly due to the increase in the ability to bind H\* on the Au(110) surface with GBs. Therefore, the GBs on the Au(110) surface can dramatically improve H\* binding strength and, thus, boost the HER process.

#### 3.1.3. Exchange Current Density

For the purpose of the application of the GB concept in the HER experiment, we calculated the exchange current density (*j*0), the most common experimental descriptor, to measure the catalysts' HER activity [10,26]. According to the micro–kinetic model, at equilibrium, *j*<sup>0</sup> can be theoretically computed as an indirect function of ΔGH\* [10],

$$\dot{y}\_0 = Fk^0 \mathbb{C}\_{total} \left[ (1 - \theta)^{1 - a} \theta^a \right] \tag{4}$$

$$\theta = \frac{\exp\left(-\Delta \mathbf{G}\_{\ast \mathbf{H}} \Big/ \Big/ \_{k\_{\mathbf{B}} \mathbf{T}}\right)}{1 + \exp\left(-\Delta \mathbf{G}\_{\ast \mathbf{H}} \Big/ \_{k\_{\mathbf{B}} \mathbf{T}}\right)} \tag{5}$$

where *k*<sup>0</sup> is the standard rate constant, *α* is the transfer coefficient (*α* was set to 0.5 in this work), *Ctotal* is the total number of HER active sites on the surface of the catalyst, and *kB* is the Boltzman constant. The relationship of catalysts correlating EB(H\*) is presented with the logarithm of *j*<sup>0</sup> (log *j*0) (Figure 3a). Similar to the UL–EB[H\*] plot, a typical volcano plot is also clearly observed, which is consistent with that in a previous study [10]. Pt and other HER–active materials are close to the top of the volcano curve, which denotes the scaling relationship between log *<sup>j</sup>*<sup>0</sup> and EB[H\*]. The flat Au(110) surface has log *<sup>j</sup>*<sup>0</sup> <sup>=</sup> −10.88 A·cm−<sup>2</sup> at EB[H\*] = 0.50 eV, which implies that Au(110) is not a good HER catalyst because its *j*<sup>0</sup> is extremely small. Once GBs are formed on the Au(110) surface, the log *j*<sup>0</sup> of the sites around the GBs is significantly increased (s1–s5), and they shift up toward the volcano top. s1–s3 are promoted several to a dozen million times in *j*<sup>0</sup> with respect to the Au(110) surface. It is more surprising that, for s1 and s2, log *<sup>j</sup>*<sup>0</sup> jumps to −3.30 A·cm−2, which is close to the volcano top, outperforming Pt(111) (log *<sup>j</sup>*0=−3.50 A·cm−2). This dramatic increase in *<sup>j</sup>*<sup>0</sup> is also ascribed to the strong H\* binding capability caused by GBs on the Au surface.

**Figure 3.** (**a**) Computational exchange current density (log (*j*0)) for hydrogen evolution over Au GB sites, and various metal surfaces plotted as a function of the hydrogen binding energy. (**b**) Current density of hydrogen evolution as a function of the applied potentials for Au(110) with varying concentrations of GBs on Au(110), and Pt(111) is presented for comparison. The inset presents a magnified version of the plot.

#### 3.1.4. GB–Mediated Current Density

In fact, the active sites on the Au surface with GBs consist of GB sites and normal sites, which both contribute to the hydrogen evolution. Thus, the overall HER performance of the Au(110) surface with GBs should be described by current density (*j*) at a given applied potential, which results from the combination of Au GBs and a flat Au(110) surface. For this purpose, we plotted the current density of hydrogen evolution as a function of the applied potential (U) using the kinetics of electrode reactions [27], which can simulate a linear scan voltammetry (LSV) curve. Under the precondition of ruling out mass transfer effects, the LSV curve of the HER can be predicted using the Butler–Volmer Equation (Supporting Information),

$$j = j\_{forward} + j\_{backward} = j\_0 \left[ e^{-af\eta} - e^{-(1-a)f\eta} \right] \tag{6}$$

where *α* is the transfer coefficient, *η* is a certain overpotential, and *f* denotes F/RT. Figure 3b shows a *j*–U plot of Au(110) surfaces with varying amounts of GBs by controlling the ratio of GB sites (s1–s3)/normal sites. The GB–free Au(110) surface has the largest overpotential of −0.73 V (vs SHE) at a current density of 20 <sup>μ</sup>A cm<sup>−</sup>2. Surprisingly, GBs induce a significant decrease in the overpotentials required for the HER. Of particular note is that only 0.01% of the GBs on Au(110) lead to a dramatic decrease in the overpotential by almost one–half from −0.73 to 0.33 (V vs. SHE), suggesting that quite a low GB density on Au(110) can extremely boost the overall HER performance. As can be clearly seen in Figure 3, the higher the density of the Au GBs, the lower the potential that the catalyst will have. When the density of the GBs increases to 1%, the overpotential to achieve 20 μA cm−<sup>2</sup> further reduces to −0.09 V (see inset). A total of 5% of GBs can further reduce the overpotential to 0.01 V. However, we cannot expect even higher HER activity by continuously increasing the GB density, because there exist only a few GBs on a real catalyst surface.

#### *3.2. Experimental Verification of GB–Assisted HER Activity*

#### 3.2.1. Exposed Surface Characterization

To determine the accessible exposed facet of the as–prepared Au electrodes, we used the lead underpotential deposition (Pb–upd) technique. The Pb–upd process is very sensitive to the surface structure of the gold electrode with a distinct voltammetric profile, where the main peaks appear at different characteristic potentials for different exposed facets of Au electrodes [28,29]. As a result, the main peaks can be used as indicators to characterize the surface structure of the electrode. In Figure 4, it can be clearly observed that the pristine Au/CFP electrode shows only one significant deposition peak in the Pb–upd voltammogram profile, which is located at ~−0.47 V (vs. Ag/AgCl) and corresponds to the (110) facet of Au. Upon scan reversal, the Pb–upd layer is oxidized to dissolved Pb2+ in the form of Pb(OH)<sup>−</sup> in alkaline solution in a sharp peak at ~−0.38 V (vs. Ag/AgCl), which is in good agreement with the results of a previous study [30]. Additionally, the CFP electrode exhibits no Pb–upd peak, suggesting that Pb–upd is attributed to the Au itself. Therefore, Au(110) is the dominant exposed surface faceting of the as–deposited Au/CFP electrodes, which provides the experimental basis for the verification of our GB–assisted HER on Au(110).

**Figure 4.** Pd−upd voltammetric profile of as−deposited Au/CFP and clean CFP electrodes (0.5 <sup>×</sup> 1 cm2). Scan rate: 50 mV/s.

We performed a detailed characterization of the morphology and structure of the as–prepared and annealed Au/CFP. From SEM and compositional maps (Figure S2), it can be seen that Au is deposited on the CFP electrode with a uniform distribution and a small particle size. The as–deposited Au/CFP was characterized by X–ray diffraction (XRD) (Figure S3). When the sputtering time is short, such as 30 s or 2 min, the intensity of the Au diffraction peaks is weak, because the content of Au is small and the particle size is small.

The morphological structures of the as–deposited and annealed Au/CFP were investigated via transmission electron microscopy (TEM) observation (Figure 5). The high– resolution TEM (HR–TEM) images show a lattice fringe spacing of 2.35 Å corresponding to the (111) facet direction of the Au nanoparticles (NPs). It can be clearly seen that the as–deposited Au NPs feature a representative surface rich in GBs, showing a polycrystalline character to some extent (Figure 5a). Notably, the grain boundary angle in Figure 5a is similar to that in our calculated GBs surface model (Figure 1).

**Figure 5.** High–revolution TEM images of as–prepared (**a**) and annealed Au/CFP electrodes at 300 ◦C (**b**) and 400 ◦C (**c**). The light yellow lines indicate GBs in the Au NPs. The white lines indicate the lattice fringe spacing for Au NPs. The scale bar represents 5 nm.

To control the density of the GBs on Au(110), we annealed the pristine Au/CFP under an Ar atmosphere at different temperatures, namely, 300 ◦C and 400 (producing Au/CFP300 and Au/CFP400, respectively). By comparing the change in the morphology before and after annealing, it is interesting to note that the (110) facet is still dominant on various Au/CFP electrodes. In addition, the density of the GBs in the Au NPs dramatically decreases as the annealing temperature is increased (Figure 5b,c). At 300 ◦C, only a few Au NPs with GBs are observed, and the GB density in the Au NPs becomes much lower, suggesting that most Au NPs are transformed into a monocrystalline form. Nevertheless, Au NPs are completely monocrystalline after annealing at 400 ◦C.

#### 3.2.2. Electrochemical HER Characterization

The effect of GBs on the HER activity of the Au NPs was examined in acid aqueous solution (0.5 M H2SO4). Figure 6a shows representative LSV curves for the CFP substrate, pristine Au/CFP and annealed Au/CFP electrodes (Au/CFP200, Au/CFP300 and

Au/CFP400). The CFP substrate has negligible HER activity and, thus, is expected to exert no influence on the current density contribution by the GBs. The pristine as–deposited Au/CFP has the highest HER activity and shows an overpotential of 330 mV at a current density of 10 mA/cm2. By contrast, the overpotentials required to achieve 10 mA/cm2 HER current density for the annealed Au/CFP electrodes are 470 mV (Au/CFP200), 530 mV (Au/CFP300) and 550 mV (Au/CFP400). In combination with the above TEM analysis, the HER activity significantly decreases with the decrease in the density of the GBs in the Au NPs. Au/CFP300 has only a 20 mV lower overpotential than Au/CFP400, as the GBs in the Au NPs almost diminish after annealing at 300 ◦C and 400 ◦C, which is consistent with the above TEM observation (Figure 5b,c). Obviously, the experimental LSV curve is in good agreement with our simulated current density curve (Figure 3b). Although it is quite difficult to measure the GB density accurately due to the irregular shape of Au NPs, the overpotential difference (220 mV) still indeed suggests that the GBs in Au NPs could significantly enhance HER activity.

**Figure 6.** Effect of GBs on Au surface on the electrochemical HER activity of Au/CFP electrodes. (**a**,**b**) Electrochemically active (EA) surface area−calibrated linear sweep voltammetry (LSV) curves and Tafel plots for four Au/CFP electrodes with a size of 0.5 <sup>×</sup> 1 cm<sup>2</sup> (electrolyte: 0.5 M H2SO4, scan rate: 5 mVs−1). (**c**) Turnover frequency per surface Au atom of various Au/CFP samples, and the inset is the TOF value obtained at an overpotential of 300 mV.

A comparison of Tafel plots for various Au/CFP electrodes is shown in Figure 6b. It can be seen that the introduction of GBs in Au NPs leads to a smaller Tafel slope compared with the GB–free counterparts, resulting in values of 145 and 185 mV dec−<sup>1</sup> for Au/CFP and Au/CFP400, respectively. As widely accepted, Tafel slopes of 120, 40 and 30 mV dec−<sup>1</sup> have been observed for the Volmer, Heyrovsky and Tafel determining rate steps, respectively [31]. Regarding the HER mechanism, the rate–determining step for this Au catalyst is the Volmer reaction, i.e., the initial adsorption of protons from the acid solution to form adsorbed H [10]. Of note is that GBs may change the rate–determining step (RDS); that is, RDS may be changed from binding H\* to desorbing hydrogen at Au atoms around the GBs by increasing the capability of binding H (Figure 2a, Table S2), which is in agreement with the theoretical ΔGH\* for the GB sites and normal sites (Figure 2a).

To exclude the improvement in HER activity caused by particle size and surface areas, we calibrated the electrochemically active (EA) surface area (Figure S4) of the as– prepared Au/CFP electrodes. Based on the EA surface area, we then calculated the turnover frequency (TOF) per surface Au atom to compare the intrinsic activity of various Au/CFP electrodes. Figure 6c shows the experimental TOF as a function of applied overpotential. The TOF value increases with the overpotential following the Tafel behavior for all Au/CFP electrodes. The TOF value of the as–deposited Au/CFP is even higher than those of the other annealed Au/CFP electrodes, indicating that, after annealing treatment, the highly effective sites for the HER significantly decrease. Additionally, the TOF values of Au/CFP300 and Au/CFP400 are almost the same over the whole overpotential range. Specifically, at an overpotential of 300 mV, the as–deposited Au/CFP electrode exhibits a high TOF value of 0.33 s<sup>−</sup>1, which is extremely higher than those of the annealed Au/CFP electrodes (0.04 s−<sup>1</sup> for Au/CFP200 and 0.01 s−<sup>1</sup> for Au/CFP300 and Au/CFP400) (see inset, Figure 6c). The trend in the TOF value for various Au/CFP electrodes is consistent with LSV and Tafel, and it is inherently correlated with the decrease in the density of the GBs on the Au surface after annealing (Figure 5). Thus, creating GBs is a much more effective strategy to improve HER activity of the originally inactive metal electrocatalyst.

#### *3.3. Origin of GB–Enhanced HER*

Although Au(110) is not a traditional active catalyst for the hydrogen evolution reaction [20], it becomes a great activity catalyst for the HER by forming grain boundaries, as stated above. The improved HER activity is ascribed to the H\* binding energy caused by the GBs in the Au electrocatalyst. As the strength to bind H\* is inherently decided by the underlying electronic structure of the catalyst itself, we focus on the differences in atomic and electronic structures, which may provide good guidance. The density of states (DOS) is calculated to understand the interaction between the H atoms and Au GB sites and, thus, to gain an insight into the different HER activities of the GB sites. Only the 5*d*–projected DOS of Au atoms is presented because Au 5*d* orbitals decide the formation of H−Au bonds. Figure 7 shows the Au 5*d* orbital–projected DOS (PDOS) onto s1, s2 and s5 on Au(110) with GBs. The 5*d*–PDOS of the flat Au(110) with H\* is also presented for comparison. Compared with the GB–free Au(110), the peak positions and shapes of Au(110) with GBs vary. The variation in the H\* binding strength for different GBs sites can be explained by a *d*–band model, in which the *d*–band center (*ε*d) is correlated with the adsorption energy of the adsorbate by considering either the adsorbed H\* state by itself or the energy–level realignment in the HER process [32]. As previously mentioned, neither a too strong nor a too weak binding strength is suitable for HER, and an appropriate *ε*<sup>d</sup> can ensure an appropriate strength of the H−Au bond and catalytic activity for the HER. The value of *ε*<sup>d</sup> can be calculated by

$$\varepsilon\_d = \frac{\int\_{-\infty}^{+\infty} E \times DOS(E) dE}{\int\_{-\infty}^{+\infty} DOS(E) dE} \tag{7}$$

Considering the dominant role of *d*–states under the Fermi level in the H\* binding process, we integrated the domain of the Au 5*d* states below the Fermi energy to obtain an effective *ε*d. As seen in Figure 7a, the *ε*<sup>d</sup> values for s1, s2 and s5 around the GBs are calculated to be −3.38 eV, −3.08 eV and −3.47 eV, respectively, higher than the GB–free Au(110) surface (−4.05 eV). That is, GBs indeed allow *ε*<sup>d</sup> to shift up toward the Fermi energy on the Au(110) surface. It is well known that a higher *d*–band center corresponds to stronger adsorption [32]. Consequently, the upper shift of *ε*<sup>d</sup> suggests that the sites near GBs have a larger H\* binding strength and, thus, a lower UL in comparison with the clean Au surface. To quantitively evaluate the relationship between HER activity and the electronic structure of Au(110) GBs, we present the trend in UL(H\*) on different GB sites as a function of *ε*<sup>d</sup> in Figure 7b. We can see that *ε*<sup>d</sup> exhibits an approximately linear relationship with UL with *r* = 0.84, suggesting that the sites just at the GBs are superior to the sites near the GBs.

**Figure 7.** (**a**) Au 5*d* orbital−PDOS on active sites on Au(110) with and without GBs. The Fermi energy (*E*f) is set to zero and is represented by a black vertical dashed line. The green, thick bar represents the *ε*<sup>d</sup> of a specific atom on Au(110) surface in the absence or presence of GBs. (**b**) The trends in UL for HER on different GBs sites as a function of CN*<sup>d</sup>* and *<sup>ε</sup>*d. The perfect Au(110) plane is also presented for comparison.

In fact, GBs feature disorder of the atomic arrangement, where under–coordinated atoms exist; we can also elucidate the origin of GB–assisted HER from a geometrical point of view. By considering the environment of their neighboring atoms, we use the concept of the bond–energy–integrated coordination number (CN*<sup>d</sup>* ) proposed by us [33,34]. CN*<sup>d</sup>* can differentiate the surface sites in HER activity. The CN*<sup>d</sup>* of the five active sites mentioned above is calculated according to the following expression [34]:

$$\overline{\text{CN}}^d = \frac{\sum\_{i \neq j}^{r\_{ij} < r\_c} \sqrt{\mu\_{id}^{(2)}}}{\sqrt{\left(V\_{nn}^{d,\infty}\right)^2}} \tag{8}$$

where *μ*(2) *id* is the second moment of the local density of states projected on the d orbital of site *i*, which is calculated by summing the square of the d–electron hopping integrals to the neighboring d orbitals by only counting the first nearest neighboring atom j within a cutoff distance of rc = 3.2 Å. *Vdsσ*, *Vddσ*, *Vdd<sup>π</sup>* and *Vdd<sup>δ</sup>* are the d–electron hopping integrals to the s or d orbital of the first nearest–neighbor atom for the formation of *<sup>σ</sup>*, *<sup>π</sup>* and *<sup>δ</sup>* bonds. *<sup>V</sup>d*,∞*nn* is the corresponding d–electron hopping integral in bulk metal, which is a constant for the catalysts of the same metal.

In Figure 7b, we can see that CN*<sup>d</sup>* exhibits an approximately linear relationship with UL, with *r* = 0.88. The sites around the GBs (s1–s5) have a smaller CN than those on the flat Au(110) surface (=7.00). Of note is that a decrease in CN*<sup>d</sup>* corresponds to a decrease in UL. As stated above, UL is strongly correlated with H\* binding strength; therefore, a smaller CN*<sup>d</sup>* implies a higher EB[H\*] for Au GBs and, thus, better HER performance. Therefore, HER activity (or H\* binding) can be directly evaluated by the CN*<sup>d</sup>* of the active sites around the GBs. Compared with the flat extended surface, atoms around GBs have no regular arrangement; that is, they are rich in atomic disorder, leading to the lack of complete coordination (unsaturated coordination). Thus, strong Au–H bonds have to form to compensate for the lack of coordination, following the bond–order conservation theory [35,36], leading to a higher H\* binding capacity of the sites around GBs. By making a comparison with the UL– CN*<sup>d</sup>* relationship, it is evident that a smaller CN*<sup>d</sup>* corresponds to a larger *ε*<sup>d</sup> shift, indicating that the lack of a coordination number alters the *ε*<sup>d</sup> of active sites around GBs. Hence, in the present case, the bond–energy–integrated coordination number or the *d*–band center theory explains the origin of the improved HER activity; both originated from the disorder of the atomic arrangement, which inherently exists at GBs.

#### **4. Conclusions**

In summary, via DFT simulations, we studied the HER performance of Au mediated by GBs. The active site around GBs exhibited superior HER performance with a remarkably low overpotential required, which arose from the significant enhanced capacity of binding H. The Au catalyst with GBs was experimentally prepared using the magnetron sputtering technique, and it exhibited superior HER activity compared to the Au catalyst without GBs. HER activity was inherently correlated with the density of the GBs in the Au catalyst, where active sites bind hydrogen more strongly. The possible reasons for this increase in H\* binding ability were discussed, including geometrical and electronic features. The origin of the GB–assisted HER was the disorder of the atomic arrangement in the GB region on the Au surface, leading to the lack of complete coordination; thus, a strong H–Au bond formed to compensate for the lack of coordination. In addition, this disorder gave rise to the migration of the *d*–band center upward to the Fermi energy, leading to increased bonding. The GB–mediated HER performance emphasizes that the GB technique may be particularly suitable for applications in practical electrochemical hydrogen production.

**Supplementary Materials:** The following supporting information can be downloaded at: https: //www.mdpi.com/article/10.3390/app12094290/s1, Figure S1: Front and side views of fully relaxed periodic supercells for ∑6{2–21} HAGB; Figure S2: SEM images of as–prepared (a) and annealed Au/CFP electrodes at 300 ◦C (c) and 400 ◦C (e). Compositional maps of as–prepared (b) and annealed Au/CFP electrodes at 300 ◦C (d) and 400 ◦C (f), with red indicating Au and green indicating carbon; Figure S3: X–ray diffraction patterns of Au/CFP for 30 s, 2 min and 5 min magnetron sputtering; Figure S4: Cyclic voltammograms under different scan rates for (a) Au/CFP, (b) Au/CFP300 and (c) Au/CFP400. (d) Electrochemically active (EA) surface area derived from a change in charging current density related to scan rate; Table S1: Calculated values for conversion of electronic energies to free energies. Assumed fugacities for gaseous H2 are also included; Table S2: Calculated values for conversion of electronic energies to free energies on the GBs on Au(110) surface. Assumed fugacities for gaseous species are also included; Table S3: Calculated values for conversion of electronic energies to free energies of metals on (111) and (211) surfaces. Assumed fugacities for gaseous species are also included; Table S4: Calculated values for conversion of electronic energies to free energies of metals on (100) and (110) surfaces. Assumed fugacities for gaseous species are also included; Table S5: The flat surface and grain boundary active sites on Au(110). The coordination numbers and band centers are given.

**Author Contributions:** Data curation, formal analysis and writing—original draft preparation, R.J.; methodology and software, J.F.; supervision, Z.W.; supervision, funding acquisition and writing review and editing, C.D. All authors have read and agreed to the published version of the manuscript.

**Funding:** This research was funded by Natural Science Foundation of China (No.21403152) and Tianjin Natural Science Foundation (16JCQNJC05700).

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

**Informed Consent Statement:** Not applicable.

**Data Availability Statement:** This study did not report any data.

**Acknowledgments:** This study was supported by the Natural Science Foundation of China (No. 21403152) and the Tianjin Natural Science Foundation (16JCQNJC05700).

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

#### **References**


## *Review* **Recent Progress in ZnO-Based Nanostructures for Photocatalytic Antimicrobial in Water Treatment: A Review**

**Ziming Xin 1, Qianqian He 1, Shuangao Wang 1, Xiaoyu Han 1, Zhongtian Fu 1, Xinxin Xu <sup>2</sup> and Xin Zhao 1,\***


**Abstract:** Advances in nanotechnology have led to the development of antimicrobial technology of nanomaterials. In recent years, photocatalytic antibacterial disinfection methods with ZnO-based nanomaterials have attracted extensive attention in the scientific community. In addition, recently widely and speedily spread viral microorganisms, such as COVID-19 and monkeypox virus, have aroused global concerns. Traditional methods of water purification and disinfection are inhibited due to the increased resistance of bacteria and viruses. Exploring new and effective antimicrobial materials and methods has important practical application value. This review is a comprehensive overview of recent progress in the following: (i) preparation methods of ZnO-based nanomaterials and comparison between methods; (ii) types of nanomaterials for photocatalytic antibacterials in water treatment; (iii) methods for studying the antimicrobial activities and (iv) mechanisms of ZnObased antibacterials. Subsequently, the use of different doping strategies to enhance the photocatalytic antibacterial properties of ZnO-based nanomaterials is also emphatically discussed. Finally, future research and practical applications of ZnO-based nanomaterials for antibacterial activity are proposed.

**Keywords:** nanomaterial; photocatalyst; antibacterial; zinc oxide; water treatment; antimicrobial

#### **1. Introduction**

With rapid global population growth, urbanization increasing, illicit misuse of freshwater resources, and continued destruction of the global climate, the increasing demand for clean water is becoming a global concern [1–3]. Seven billion people, more than 15% of the world's people, are facing a shortage of fresh water resources, which even causes them to not have enough fresh water to sustain normal life and productive work [4,5]. Water scarcity is exacerbated by the increasing water pollution from releases of waterborne pathogens, inorganic pollutants, organic pollutants, agricultural chemicals, derivatives of human and animal drugs, and endocrine disruptors [6–8].

Infectious diseases caused by biological contamination such as typhoid fever, dysentery, cholera, and diarrhea are a major cause of death worldwide and continue to replicate at an alarming rate [9]. The extensive use of antibiotics and antibacterial drugs has led to strong drug resistance in viruses and bacteria, which further exacerbates the spread of biological infectious diseases [10]. In addition, the recent epidemics of global security issues such as the COVID-19 virus and monkeypox virus are caused by the spread of viruses that threaten all human beings and have a great impact on human production, life, and health [11–14]. Similar to bacteria and pathogenic microorganisms, epidemic viruses are always difficult to eradicate due to the abuse of antibiotics and various disadvantages of disinfectants [15,16]. In view of the above situation, it is crucial to explore more effective solutions and approaches.

Recent advancements in semiconductor materials and new nanomaterials have blazed new trails for their applications in the fields of photocatalysis and bacterial inactivation [9].

**Citation:** Xin, Z.; He, Q.; Wang, S.; Han, X.; Fu, Z.; Xu, X.; Zhao, X. Recent Progress in ZnO-Based Nanostructures for Photocatalytic Antimicrobial in Water Treatment: A Review. *Appl. Sci.* **2022**, *12*, 7910. https://doi.org/10.3390/ app12157910

Academic Editor: Anna Annibaldi

Received: 15 July 2022 Accepted: 5 August 2022 Published: 7 August 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/).

Nanotechnology provides a variety of promising nanomaterials for the field of photocatalytic antimicrobial. Metal oxides have many advantages such as non-toxic, stable, and efficient biological properties, which make them stand out among many nanomaterials and become a research hotspot in this field. Numerous nanomaterials doped with metal oxides such as ZnO [17], Fe2O3 [18], TiO2 [19], Ag2O [20], CaO [20], MgO [21], and CuO [22] have been applied as efficient antibacterial agents for both Gram-positive (G+) and Gramnegative (G-) bacteria, such as *Escherichia coli*, *Salmonella enteritidis*, *Streptococcus pyogenes*, *Aeromonas hydrophila*, *Pseudomonas aeruginosa*, *Salmonella typhimurium*, *Fecal intestinal cocci*, etc. [23]. Among the antibacterial agents, ZnO-based nanomaterials are widely recognized as promising antibacterial agents with strong photocatalytic antibacterial activity [24,25]. Nevertheless, the wide bandgap of ZnO is approximately 3.2–3.3 eV, which affects its light absorption ability, resulting in a response only in the ultraviolet band [26]. Previous studies indicated that defects of nanomaterials in photocatalytic antibacterial processes could be effectively improved after being modified [27]. Strategies, such as loading antibacterial agents, loading oxidized nanomaterials, and adjusting the particle size, material microshape, and concentration of ZnO were employed to enhance the antibacterial properties. While exploring the antibacterial ability improvement, the antibacterial mechanism should also be in-depth investigated.

In previous studies, some mechanisms for photocatalytic antimicrobials, such as metal ion release, reactive oxygen species (ROS) generation [28], destruction of cell membranes, internalization of nanoparticles [29], interruption or blockade of transmembrane transport, etc., have been proposed [30,31]. Overall, the purpose of all antibacterial mechanisms is to disrupt the bacterial cell structure and break it down into harmless substances. However, our understanding of the specific process of substance transformation during the photocatalyst-induced antimicrobial process is still very limited, which requires further exploration.

This review is an exhaustive summary of recent research advances on the antimicrobial of ZnO-based nanomaterials. By summarizing previous studies, ZnO-based nanomaterials with excellent antibacterial effects were obtained. From the perspective of material preparation, different preparation methods are reviewed, and both the advantages and disadvantages are compared. Subsequently, the antimicrobial mechanism of ZnO-based nanomaterials is discussed in-depth from both the physical and chemical aspects. In detail, various strategies to enhance the antimicrobial ability of ZnO-based nanomaterials in recent studies are proposed. Finally, temporary deficiencies in the improvement strategy are summed up, and prospects for the future development direction and application potential are presented.

#### **2. ZnO-Based Nanostructures Preparation**

Among the numerous methods for preparing ZnO-based nanomaterials, the wetchemical/solution technique has many advantages such as simplicity, rapid operation, and cost savings, which make it a promising method for the preparation of ZnO-based nanomaterials. The advantages and disadvantages of commonly used wet-chemical/solution techniques, such as the sol–gel method, co-precipitation method, microwave-assisted method, and hydrothermal method, are presented in detail in Table 1. The preparatory stage of ZnO nanomaterial growth is fully wetted by wet-chemical/solution techniques, which greatly improves the stability of the materials.


**Table 1.** Advantages and disadvantages of different antibacterial synthesis methods.

#### *2.1. Sol–Gel Method*

The sol–gel method is one of the most effective chemical methods for nanocomposites preparation with desired properties and advantages, such as low cost, mild reaction, environmental friendliness, reliability, and simplicity [43]. In the preparation of photocatalytic materials, the materials synthesized by this method have better photocatalytic activity [44]. ZnO nanoparticles were successfully synthesized using the gel-sol method by Hasnidawani [45], and the surface morphology was verified by Fe-SEM images (Figure 1), which was confirmed to have a rod-like structure with a dense particle structure. Varieties of ZnO nanostructures have been discovered, which are in the form of nanorods, nanotubes, nanobelts, nano springs, nano spirals, nano rings, and many more [46]. Among these structures, the rod-like structure is the best nanostructure compared to others due to their one-dimensional nanostructures (such as nanorods, nanowires, and nanotubes) that can facilitate more efficient carrier transport for the decreased grain boundaries, surface defects, disorders, and discontinuous interfaces [47,48]. In the process of preparing ZnO-based nanomaterials by the sol–gel method, the influence of factors such as solution drop acceleration rate, reaction temperature, pH, etc., will have a significant impact on the antibacterial properties of the materials [49]. Effects of various preparation influencing factors on the antibacterial properties of ZnO were tested, and pH was proven to be the most important influencing factor [50]. The reason is suggested to be that the neutral and acidic solution environment is more suitable for Zn2+ to function and achieve an antibacterial effect. As long as the optimal pH and preparation temperature are found, the sol–gel method will be one of the effective preparation methods with high efficiency and low cost. Therefore, in recent studies, the sol–gel method is used more in the synthetization of ZnO-based nanoparticles [51–53].

**Figure 1.** FE-SEM micrographs of synthesized ZnO at different magnifications. Reprinted/adapted with permission from Ref. [45]. Copyright © 2022, Elsevier B.V.

#### *2.2. Co-Precipitation Method*

The co-precipitation method does not require expensive raw materials and complicated equipment, which provides a suitable method for low-cost and large-scale production [54]. Furthermore, in addition to simpler devices, suitable metals, metal oxides, and surfactants are added to change the morphology of the materials [55]. The co-precipitation method was chosen to prepare ZnO-based nanoparticles in an aqueous solution at two different reaction temperatures (50 ◦C and 70 ◦C) by Kotresh et al. [56]. The surface morphology of ZnO nanoparticles prepared by the co-precipitation method was observed by scanning electron microscope (SEM) images as shown in Figure 2. It can be seen from the SEM images that the particles are uniformly spherical with a dense and dense structure. The spherical ZnO nanoparticles prepared by the co-precipitation method are favorable for uniform dispersion in the photocatalytic reaction and efficiency improvement of photocatalytic reactions. However, it was found that the droplet acceleration rate had the greatest impact on the antibacterial properties of ZnO-based nanomaterials synthesized by the co-precipitation method [57]. The size of the nanoparticles synthesized by the co-precipitation method is affected by the drop rate of the solution, which will affect the contact area of the nanoparticles during the antibacterial reaction, thereby greatly affecting the antibacterial effect [58]. Therefore, the droplet acceleration rate and particle size need to be carefully considered during synthesis, which is an important part of the success of the co-precipitation method. Due to the advantages of the co-precipitation method with a simple preparation process, the use of this method has gradually increased in nanomaterial synthesis research in recent years [59–61].

**Figure 2.** SEM images of sample A and sample B. Reprinted/adapted with permission from Ref. [56]. Copyright © 2022, Elsevier GmbH.

#### *2.3. Microwave-Assisted Method*

The microwave-assisted method is not only an energy-saving, environmentally friendly, and heat-free method, but also with many advantages such as fast synthesis speed and the ability to tune the particle shape [44,62]. ZnO nanoparticles with different morphologies can be synthesized by the microwave-assisted hydrothermal method by adjusting the time and power of microwave irradiation [63,64]. Through microwave-assisted chemistry techniques, ZnO nanostructures with different morphologies were synthesized in different pH reaction mixtures (acidic, basic, or neutral). Furthermore, nanomaterials are synthesized without any heating and addition of surfactants. Hence, obtaining ZnO particles with oxygen vacancies and defects is expected to improve their pollutant degradation behavior due to the fast reaction process and non-stoichiometric synthesis [65]. As shown in Figure 3, the microscopic morphology of the microwave-synthesized ZnO nanostructures was observed by SEM. It can be seen intuitively that the appearance of ZnO nanoparticles changes dramatically with the pH change. In addition, from the XRD analysis in Figure 4, it was demonstrated that the change in intensity and peak width of the two sets of samples (prepared with NaOH and KOH as pH control agents) can be observed as the solution pH changes. The above results showed that the shape of synthesized ZnO nanoparticles is affected by pH changes, which has important implications for the directional synthesis of nanoparticles with diverse morphologies. In addition, several studies have shown that the power of the microwave is the most important factor affecting the synthesis of ZnO-based nanomaterials by microwave-assisted method [66,67]. Moderate-power microwave have been shown to be suitable for the synthesis of ZnO-based nanomaterials with stronger antibacterial capabilities [68,69]. However, the exact power influence mechanism needs to be further explored.

**Figure 3.** Typical SEM images of synthesized ZnO nanostructures at different reaction pH values. Reprinted/adapted with permission from Ref. [65]. Copyright © 2022, Elsevier Ltd and Techna Group S.r.l.

**Figure 4.** Typical XRD patterns of ZnO nanostructures synthesized with different reaction pH ((**a**) NaOH and (**b**) KOH as pH controlling agents, 7, 8, 10, 12 correspond to their respective concentrations). Reprinted/adapted with permission from Ref. [65]. Copyright © 2022, Elsevier Ltd. and Techna Group S.r.l.

#### *2.4. Hydrothermal Method*

The hydrothermal method is a convenient and highly efficient method, which requires a lower reaction temperature and saves costs [70]. In addition, by adjusting the duration, density, and reaction temperature of the contained substances, the morphology and size of the particles can be controlled [71]. As shown in Figure 5a–c, TEM images of synthesized ZnO particles at precursor concentrations of 5, 10, and 20 mM are revealed. The morphological features of the poorly dispersed nano-ZnO crystals are clearly demonstrated by the TEM images in all cases. At precursor concentrations of 5, 10, and 20 mM, ZnO nanoparticles were observed to have diameters of approximately 4.5, 6, and 8–9 nm, respectively. Furthermore, it was also observed from the images that the size distribution of the nanoparticles was fairly uniform [72]. As shown in the XRD pattern (Figure 6), the crystalline structure of the synthesized ZnO particles after hydrothermal treatment was confirmed. Simultaneously, no impurity peaks were detected from the XRD pattern, indicating that the target substance with a higher purity was successfully synthesized [72]. Furthermore, the preferred orientation of the ZnO particle samples was not seen from XRD patterns, suggesting that ZnO crystals may have the most shapes other than rods or sheets. Based on the above conclusions, the hydrothermal method is a suitable method to prepare ZnO-based nanoparticles with different shapes. In addition, the antibacterial effect of ZnO-based nanomaterials prepared by hydrothermal method is affected by several factors, such as pH, reaction temperature, and dosage ratio [73,74]. The reaction temperature directly affects the structure of the material and changes the antibacterial ability, while the pH changes the surface properties and shape of the material to affect the antibacterial ability [75].

**Figure 5.** Electron microscope image of ZnO nanoparticles: the concentration of precursor is (**a**) 5 mM, (**b**) 10 mM, and (**c**) 20 mM. Reprinted/adapted with permission from Ref. [72]. Copyright © 2022, Elsevier Ltd.

**Figure 6.** XRD patterns of ZnO samples synthesized with different precursor concentrations. Reprinted/adapted with permission from Ref. [72]. Copyright © 2022, Elsevier Ltd.

#### **3. Types of Nanomaterials for Photocatalytic Antimicrobials in Water Treatment**

A variety of nanomaterials as efficient adsorption materials and catalytic degradation and purification of wastewater will be discussed in this section. As shown in Figure 7, various microbial contaminants and their sources, as well as different nanomaterials for removal applications are revealed [76]. Limitations of single nanomaterials can be overcome

by various strategies such as polymer/metal oxide nanocomposite [77], metal oxide/metalbased nanomaterials [78,79], polymer/metal oxide nanocomposite [80,81], and polymerstructure-based materials [82,83].

**Figure 7.** Schematic representation of microbial contamination sources and their various nanomaterials for water treatment. Reprinted/adapted with permission from Ref. [76]. Copyright © 2022, Elsevier Ltd.

The nanocomposite is a safe, non-toxic, green and environmentally friendly nanomaterial, usually prepared with ZnO and TiO2 as substrates. The addition of TiO2 and ZnO NPs in some polymers such as polypropylene matrix would increase the dielectric constant of the nanocomposite, thereby enhancing the photocatalytic ability of the material [84]. In addition, the polymer boundary layer transition zone forms a crystalline structure, which increases conductivity and acts as a tuning surfactant [85], which greatly enhances the photoreactivity of the catalytic material. A variety of composite nanomaterials have been reported that can be applied to organic pollutants and microbial contamination removal from water [86]. Moreover, the addition of Ag, zinc, zeolite, and titanium is obtained having better efficiencies in pathogenic pathogens and microorganisms removal from water [87,88]. For example, silver NPs with polyurethane will flexibly remove almost 100% of *B*. *subtilis* and *E*. *coli* from water [89,90].

Among the composite nanomaterial antibacterial agents, ZnO-based nanomaterials have the advantages of strong compatibility, green friendliness, lower costs, and simple preparation, so they occupy a large proportion of the related studies in this field in recent years. Studies showed that ZnO has stronger direct interactions with bacterial and pathogen cell surfaces than other semiconductor materials [91]. In addition, ZnO nanomaterials leak Zn2+ in solution, which can exacerbate toxicity to bacteria, pathogens, and viruses. Compared with Cu+, Fe2+, and Al3+, Zn2+ showed a better ability to fight microbial contaminants [92]. Therefore, ZnO-based nanomaterials have a better potential to combat microbial contamination, which is of practical significance for the study of antimicrobial contamination. Among types of microbial pollution in the water environment, bacterial pollution is still the most important problem to be solved. Thus, in the next section, the antimicrobial properties and mechanisms of ZnO-based nanomaterials, as well as methods for studying antimicrobial properties are introduced, and the methods for improving their catalytic properties are summarized and discussed.

#### **4. ZnO-Based Nanomaterials for Antimicrobial Application in Water Treatment**

The antibacterial activity of ZnO-based nanomaterials is greatly affected by the morphology and particle diameter. Considerable methods for the preparation of ZnO-based

nanomaterials with different morphologies have been reported in the literature, such as nanorods [93–96], nano/micro flowers [97–101], microspheres [102], nano powders [103–105], nanotubes [4], quantum dots [106,107], films [107], nanoparticles [108], and capped nanoparticles [109], to understand their application prospects for antibacterial agents. To conduct an in-depth exploration of the antibacterial properties of ZnO-based nanomaterials, the antibacterial research methods in this chapter are firstly introduced, and then the antibacterial mechanism and material improvement strategies are expounded to discuss the latest studies in this field.

#### *4.1. Research Methods for Antimicrobial Activities of ZnO Nanostructures*

To better study the effect of antibacterial agents, many techniques have been adopted to test antibacterial properties in recent years. As shown in Table 2, various kinds of adsorbents for microbes removal in different water resources are listed, which has great reference significance for antibacterial application studies. In the process of antibacterial ability testing, the accuracy of the results is affected by many factors, such as types of bacterial species and the type of data to be tested, etc. In addition, external factors such as the experimental environment also affect antibacterial detection. Commonly, the temperature of 37 ◦C and incubation time of 24 h are selected for the tests. Moreover, culture media of Tryptic Soy Broth, Luria–Bertani (LB) broth, Nutrient Agar, and Tryptic Soy Agar (TSA) are commonly used. Meanwhile, depending on the technique used, parameters such as minimum inhibitory concentration (MIC), zone of inhibition (ZOI), colony count, or optical density of bacterial cultures are selected as the basis for the evaluation. The specific application methods and practical application cases of various antibacterial detection technologies are introduced below, which are beneficial for better understanding the antibacterial properties of ZnO-based nanomaterials.


**Table 2.** Various kinds of adsorbents for microbial removal in different water resources.

#### 4.1.1. Disk-Diffusion Method

The disk-diffusion method is a simple and efficient antimicrobial test, also known as the Kirby–Bauer antibiotic test (KB test) [133]. Mueller–Hinton agar and Brucella blood agar are commonly used as media for this method [133]. During the disk-diffusion method, the pH of the medium is usually controlled at around 7.2. The bacterial suspension containing a specific concentration was spread on the above agar medium and the experimental environment was required to be absolutely dry [22,24,93,94,97,134–136]. Under sterile conditions, a certain number of ZnO-based nanomaterials were soaked on the filter paper disk using the selected solvent according to the specific requirements of the experiment. Let the disk dry and carefully mount it on the medium in a Petri dish. Solvent-soaked disks were used as controls to ensure the accuracy of the test. Next, the dishes were incubated at 37 ◦C for 24 h, providing the right conditions for bacterial growth. Subsequently, due to the bactericidal activity of the ZnO-based nanomaterials, no bacterial growth was observed around the disks at specific distances. The minimum concentration at which ZnO-based exhibited antibacterial activity was called the minimum inhibitory concentration (MIC) and the area around the disk with no bacterial growth observed was called the zone of inhibition. The lower the MIC value, the larger diameter of the inhibition zone and the higher antibacterial activity. Based on this, the antibacterial properties of ZnO-based nanomaterials were judged [137,138].

#### 4.1.2. Well-Diffusion Method

The medium used in the well-diffusion method is similar to the method described in Section 4.1.1, and the two methods can be used together analogously. In contrast to the diskdiffusion method, the filter paper disk is installed by drilling holes in the media plate. The wells were filled with various concentrations of ZnO nanoparticle suspensions for testing. In addition, the calculation method of MIC and ZOI can also refer to the disk-diffusion method [108,139–142]. Consistent with the disk-diffusion method, sterile conditions are also one of the most necessary environmental conditions for antimicrobial performance testing. Based on MIC and ZOI, the pros and cons of antibacterial properties of ZnO-based nanomaterials can be studied and the efficiency can be evaluated.

#### 4.1.3. Antimicrobial Measurements in Liquid Culture Media

During the incubation period, the turbidity of the growth solution increases with bacterial growth. The liquid turbidity and cell proliferation can be measured by periodically measuring the optical density. The technique does not require reagents and special handling [143–147]. The basic principle of this method is to judge the quality of antibacterial performance by observing the absorbance at a specific wavelength in a spectrophotometer and regularly monitoring the corresponding bacterial growth. Umar et al. used ZnO nanomaterials as antibacterial agents to conduct growth inhibition experiments on *E. coli* and achieved excellent experimental results as shown in Figure 8 [101].

**Figure 8.** Bacterial growth curve of *E*. *coli* under different ZnO-NFs concentrations.

#### 4.1.4. Colony Unit Measurements

Colony unit measurement, also known as the diffusion plate technique, is often used to count the amount of living bacteria cells. ZnO-based nanomaterials are introduced into agar plates or dispersed as suspensions in specific liquid media. The strains were mounted on agar plates and incubated at 37 ◦C for a specific time. Colony forming units are counted using an appropriate counting method. In addition, the colony forming unit (CFU) value can be used to judge the antibacterial ability of ZnO-based nanomaterials [148–151]. Stankovic et al. [152] calculated the percentage of bacterial cell reduction (R%) using Equation (1).

$$\text{R\%} = \frac{\text{CFU}\_{control} - \text{CFU}\_{sample}}{\text{CFU}\_{sample}} \tag{1}$$

where *CFUcontrol* = numbers of *CFUs* per milliliter for the negative control, and *CFUsample* = *CFUs* per milliliter in the presence of ZnO dispersion.

Within a certain range, the above formula can be used to calculate the bacterial concentration, and the antibacterial performance can be investigated based on the calculation results.

#### 4.1.5. Microtiter Plate Method

The microtiter plate method, also known as the microplate method, is a method of performing antimicrobial testing by observing changes in a variable number of small test tubes or plates of microwells. Resazurin [153], 2,3,5-triphenyltetrazolium chloride (TTC), crystal violet [4] and 3-(4,5-dimethylthiazol-2-yl)-2,5-diphenyltetra-zolium bromide (MTT) [154], etc., are put into the wells as indicator solutions. Next, known or different concentrations of the test strain was dispersed into the wells of a microtiter plate. A known concentration of ZnO nanomaterials was then dispersed into the wells after dilution in a sterile broth medium. The palate was then incubated at 37 ◦C for timed intervals. The bacterial cell activity can be judged by the change of absorbance to detect the antibacterial ability.

#### *4.2. Mechanisms of ZnO-Based Antimicrobial Nanomaterials*

#### 4.2.1. Mechanisms of ZnO-Based Photocatalytic

In photocatalysis, an electron–hole couple is created under light force by reduction or oxidation reactions on the catalyst surface. The photocatalytic degradation mechanism of ZnO-based photocatalyst on pollutants is shown in Figure 9. Photocatalysis occurs when a ZnO-based photocatalyst is illuminated by light with energy greater than its band gap energy [27]. Charge separation is triggered by a light energy absorption process, which excites electrons from VB to CB, leaving holes in VB [155]. Subsequently, the photogenerated e−/h+ supports a move to the ZnO photocatalyst surface. Simultaneously, e− and h+ recombine, which reduces the quantum yield. The level of this recombination rate is affected by many factors, such as the structure of the photocatalyst and the surface modification process of the photocatalysts [156,157]. The ZnO surface is aggregated with reactive e− and h+, which promote oxidation and reduction reactions that generate excess ROS, including superoxide anion (·O2 <sup>−</sup>) hydroxyl radicals (·OH). Furthermore, the redox potential of the CB bottom of ZnO is more negative than that of O2/O2 −. Therefore, these excited electrons can generate O2 −·. Simultaneously, the top of the VB of ZnO is more positive than the redox potential of ·OH/H2O. Consequently, H2O molecules can be oxidized by these holes to form hydroxyl radicals. These highly reactive radicals (·OH, O2 −·) directly oxidize organic pollutant molecules in solutions.

**Figure 9.** Basic mechanism of ZnO photocatalysis. Reprinted/adapted with permission from Ref. [27]. Copyright© 2022, Elsevier B.V.

4.2.2. Chemical Effect of ZnO-Based Nanomaterials on Antibacterial

In the process of exploring the antibacterial mechanism of ZnO-based nanomaterials, three main chemical antibacterial mechanisms were obtained: generation of reactive oxygenated species (ROS) [158], release of Zn2+ ions [31], and photoinduced production of H2O2 [159]. The ROS mechanism is basically the same as the photocatalytic mechanism mentioned in Section 4.2.1.

Similar exhaustive studies carried out by Li et al. [31] and Song et al. [160] demonstrated that the toxicity of Zn2+ ions to cells is one of the mechanisms for the sterilization of ZnO-based nanomaterials. The concentration of ZnO NPs in ultrapure water in the toxicity test was 5 mg/L. To compare the cytotoxicity, a Zn2+ ion solution also prepared in ultrapure water was used (concentration below 0.1 mg/L). TEM images of the treated G− strains of *E*. *coli* are shown in Figure 10. It is evident that the morphology of *E*. *coli* was deformed after modification with ZnO-based nanomaterials or Zn2+ ion solution. Figure 10b,c shows intracellular fluid leakage due to Zn2+ ions and osmotic stress. The experimental results showed that the cytotoxic effects of ZnO NPs and Zn2+ ion-treated solutions on *E*. *coli* were comparable. This fully proves that the release of Zn2+ has a positive effect on antibacterial, and it also shows that it is one of the antibacterial mechanisms of ZnO-based nanomaterials (Figure 11).

**Figure 10.** TEM images of (**a**) untreated *E. coli* cells, (**b**) treated with ZnO nanoparticles, and (**c**) treated with Zn2+ ions solution in ultrapure water. Reprinted/adapted with permission from Ref. [31]. Copyright © 2022, American Chemical Society.

**Figure 11.** Schematic diagram of cell damage to G+ bacteria by Zn2+ ions. Reprinted/adapted with permission from Ref. [27]. Copyright© 2022 Elsevier B.V.

Besides the light-induced reactive oxygen species produced by ZnO NPs, many pieces of literature consider H2O2 as the main substance for antibacterial activity against dermo bacteria [28,161]. Negatively charged ROS cannot penetrate bacterial cell walls, but H2O2 can also easily penetrate bacterial cell walls. Sawai et al. [162] believed that the H2O2 produced by ZnO slurry was the main reason for the biocidal mechanism. It can also be hypothesized that after H2O2 or HO disrupts the membrane, ROS can penetrate the cell wall and enter the intracellular space, thereby enhancing the biocidal effect (Figure 12).

**Figure 12.** Schematic diagram of the damage to bacterial cells caused by ZnO nanomaterials producing H2O2. Reprinted/adapted with permission from Ref. [27]. Copyright© 2022, Elsevier B.V.

4.2.3. Influence of Physical Effects of ZnO-Based Nanomaterials on Antibacterial Performance

In the process of exploring the antibacterial mechanism of ZnO-based nanomaterials, there are three main chemical antibacterial mechanisms: plasma membrane disruption through ZnO interactions [163], cellular internalization of ZnO-based nanoparticles [164], and mechanical damage of the cell envelope [165].

At suitable pH, the bacterial surface is negatively charged due to the dissociation of carboxyl and other functional groups. Meanwhile, ZnO-based nanomaterials are positively charged with a zeta potential of +24 mV [161]. As shown in Figure 13, the opposite charges carried by bacterial cells and ZnO NPs are the reasons for the strong electrostatic attraction between them. Strong electrostatic interactions force particles larger than 10 nm in size to accumulate on the outer surface of the plasma membrane and neutralize the surface potential of the bacterial membrane, resulting in increased surface tension and membrane depolarization. In addition, strong electrostatic interactions can induce bacterial cell changes such as changes in cell membrane and membrane vesicle structure, rupture, morphological changes, and components, leading to bacterial cell death [166,167]. Since interactions play an important role in the bactericidal effect, surface modifiers and templates of ZnO-based nanomaterials would enhance the interaction with bacterial cell walls.

**Figure 13.** Schematic diagram of the physical action of ZnO nanomaterials for sterilization. Reprinted/adapted with permission from Ref. [27]. Copyright© 2022, Elsevier B.V.

Another important mechanism is cellular internalization. Cellular internalization is simply summarized as those nanostructures with a size of less than 10 nm pass through the plasma membrane, accumulate in bacterial cells, and destroy intracellular components such as nucleic acids [164,168–170]. In addition, it has also been suggested that the cellular interaction of ZnO with bacteria can enhance cell permeability (Figure 14) [171]. In conclusion, cellular internalization is one of the physical methods that plays an important role in the antibacterial process of ZnO-based NPs.

**Figure 14.** TEM images of *B*. *atrophaeus* (**a**) control, (**b**) ZnO powders, (**c**) ZnO nanorods, and (**d**) ZnO nanoparticles. Reprinted/adapted with permission from Ref. [172]. Copyright © 2022, Elsevier B.V.

The last physical mechanism is to use ZnO NPs to create cell membrane damage to destroy bacterial cells and achieve antibacterial effects. Compared with bulk ZnO materials, the presence of surface defects, uneven surface texture, and rough edges and corners on the surface of ZnO-based nanomaterials can lead to effective abrasiveness, resulting in excessive mechanical damage to bacterial cell membranes [163].

#### *4.3. Effects of Radiation Types on the Antibacterial Activity of ZnO-Based Nanomaterials*

In the process of photocatalytic antibacterial studies, ultraviolet (UV) light, sunlight, and other visible light are the most common types of radiation. In previous studies, ZnO nanomaterials have always been used for antibacterial testing under UV light due to their high band constraints [173–175]. From the work of Joe et al. [174], an important conclusion was found that the oxygen vacancy of ZnO crystals enhanced the photogeneration of ROS, and ZnO nanoparticles (NPs) with polar facets exhibited the most significant effect of antibacterial activity under UV light stimulation. Furthermore, Ma et al. creatively combined N-halamine-based materials with ZnO to improve the stability of ZnO's antibacterial performance under UV light. As a simple and effective method for nanoparticle modification, this technique can be further extended to the application of ZnO nanoparticles in other polar substrates for antibacterial functionalization [176].

Despite the promising antibacterial performance of ZnO-based nanomaterials under UV-driven radiation, photocatalysis using UV-active semiconductors is difficult due to the limited use of the solar spectrum. Simultaneously, significant progress has been made in photocatalysis using visible-light-active heteronanostructured semiconductors due to their simplicity of use, practicality, reproducibility, reliability, and commercialization [177,178]. Recently, several studies have found that photocatalytic efficiency can be improved by promoting the surface charge transfer reaction of ZnO. In addition, it may also affect the absorption spectrum of many metal oxide nanoparticles including ZnO, which enables the composites to undergo photocatalytic reactions under visible light [179–183]. The antibacterial activity of ZnO-based nanomaterials driven by visible light is of great significance for the development of photocatalytic antibacterials. Compared with ultraviolet light, visible light is more accessible, which makes visible light photocatalysis one of the hottest research fields. In addition, studies have shown that the ability of ZnO-based nanomaterials to excite ROS under visible light will be greatly improved, which also leads to the improvement of antibacterial ability under visible light [184,185]. Therefore, the design and preparation of visible-light-driven ZnO-based nanomaterials will become a meaningful research direction in future explorations.

#### *4.4. Strategies for Enhancing ZnO-Based Nanomaterials Antibacterial Activity* 4.4.1. Alkaline Earth Metal Doping into ZnO

Common alkaline earth metals including Ca, Mg, Al, and Sr, are frequently used to introduce significantly altered NPs, such as lattice defects and ionic radius differences between metal ions and Zn2+ ions, which can improve optical properties and photodegradation activity of catalysts. Taking Sr as an example, it played an important role in enhancing the catalytic degradation ability of commonly used metal oxides such as ZnO, TiO2, which can be used for photocatalytic degradation of organic pollutants in wastewater and photocatalytic antibacterial [186]. Due to the lack of local d orbitals in alkaline earth metals and the presence of local d orbitals in transition metals, the doping of alkaline earth metals is more effective in reducing the optical threshold energy of semiconductors than that of transition metals [187].

Antibacterial activity of Mg-doped ZnO nanostructures was investigated by Okeke et al. [188] towards *E*. *coli*, *P*. *aeruginosa* and *Staphylococcus aureus*. The zones of inhibition diameter of the Zn1−xMgxO sample against the selected bacteria pathogen are displayed in Figure 15. The results showed that all samples were susceptible to bacteria. The presence of more reactive sites allows surface defects to create space for ZnO to interact with microorganisms [189]. Therefore, the increase of surface defects in ZnO nanostructures

increases the reaction sites and the rate of interaction with microorganisms. Bacteria are 250 times larger than nanoparticles [105], while the bacterias have a larger relative surface area, which makes it easier for nanoparticles of much smaller size to enter the interior of bacterial cells and cause damage.

**Figure 15.** (**a**–**c**) Diameter of inhibition zones of Zn1−xMgxO against bacteria pathogen. Reprinted/adapted with permission from Ref. [188]. Copyright© 2022, Elsevier Ltd.

4.4.2. Transition Metals Doping into ZnO

The addition of transition metal ions can generate electronic states in the intermediate bandgap region to change charge separation and recombination kinetics, which is beneficial to enhancing the ability of photocatalytic antibacterial [190]. Numerous common transition metals, such as Co, Cu, Ni, Fe, and Mn, are often doped into ZnO to enhance its photocatalytic antibacterial ability [191].

Fe is the most used metal for enhancing the photocatalytic antibacterial ability of ZnO. Iron has two oxidation states with ionic radii of 0.61 Å and 0.55 Å, which are used as dopants for zinc lattice sites, respectively, since their ionic radii are smaller than those of zinc +2 (0.74 Å) radius. Hence, the doping can be alternative or interstitial to increase the conductivity of the product. Chandramouli et al. [192] doped Fe into ZnO and investigated the antibacterial properties of the composites. As shown in Figure 16, TEM images of Figure 16a pure, Figure 16b Fe doped, and Figure 16c capped ZnO NPs are observed. For undoped and doped ZnO NPs, they are more spherical with dimensions of 17–19 nm. Simultaneously, the agglomeration phenomenon occurs when ZnO is doped with Fe, which is due to the greater surface area and energy. Furthermore, the antibacterial activity against *E*. *coli* showed that the capped ZnO NPs were less toxic to the organism than ZnO NPs. As shown in Figure 17, the XRD patterns associated with various different ZnObased nanomaterials are displayed. Iron doping of ZnO reduces the grain size, resulting in a further increase in the grain size of glucose-terminated ZnO nanoparticles, which is consistent with previous reports [193]. Therefore, Fe-ZnO nanoparticles have good antibacterial activity.

**Figure 16.** TEM images of (**a**) pure, (**b**) Fe doped, and (**c**) capped ZnO nanoparticles. Reprinted/adapted with permission from Ref. [192]. Copyright© 2022, Elsevier B.V.

**Figure 17.** XRD patterns of various ZnO-based nanomaterials. Reprinted/adapted with permission from Ref. [192]. Copyright© 2022, Elsevier B.V.

#### 4.4.3. Noble Metals Doping into ZnO

Due to the formation of Schottky barriers at the metal–semiconductor interface, noble metal ions (such as Au, Ag, Pd, etc.) doped on the ZnO surface are considered to be excellent photogenerated electron traps. In addition, noble metals delay electron–hole recombination by preventing photoexcited electrons from returning to the ZnO surface, which greatly enhances the photocatalytic antibacterial ability of the composites [194]. Of all the precious metals, silver is the most stable and suitable dopant with good thermal conductivity and electrical conductivity, which will better play the photocatalytic effect of the composite material. Therefore, it has potential as a catalyst. The surface plasmon resonance (SPR) properties of silver also contribute to visible light absorption and subsequent electron–hole pair generation for the degradation of contaminants in water and for antibacterial [195].

Ye et al. [184] reported the synthesis of a series of ZnO/Ag2MoO4/Ag(ZAA) samples with theoretical molar ratios of ZnO and Ag2MoO4 of 20:1, 30:1, and 60:1 by ultrasonicassisted hydrothermal synthesis, and named the corresponding products as ZAA-20, ZAA-30, and ZAA-60 to investigate the optimal Ag2MoO4/Ag loadings. As shown in Figure 18, the characteristic diffraction peaks of ZnO and Ag2Mo4 can be clearly found in the XRD patterns, which indicates that the ZnO/Ag2Mo4 composite was successfully synthesized. The sharp diffraction peaks reveal the ultra-high crystallinity of the ZnO-based nanocomposites. In addition, with the increase of the molar ratio of ZnO to Ag2MoO4, the diffraction peak intensity of ZnO on the (002) and (110) crystal planes gradually weakened, while the diffraction peak intensity of Ag2MoO4 gradually increased. The antibacterial properties of ZnO nanosheets and ZAA nanocomposites against different contents of G− *E*. *coli* and G+ *S*. *aureus* were evaluated by the visible light electroplating counting method. In Figure 19, the bacterial cell numbers of all nanocomposites were shown to decrease with increasing contact time, and the photocatalytic antibacterial activities of the four ternary ZAA nanocomposites were much better than that of pure ZnO sheets. The experimental results demonstrate that the addition of noble metal Ag will significantly improve the antibacterial properties of ZnO nanomaterials.

**Figure 18.** XRD patterns of ZnO, Ag2MoO4/Ag, ZA, and ZAA nanocomposite with different proportions. Reprinted/adapted with permission from Ref. [184]. Copyright© 2022, Elsevier B.V.

**Figure 19.** Photographs of antibacterial test results of ZnO, Ag2MoO4/Ag, and ZAA samples against Escherichia coli (**a**–**c**) and Staphylococcus aureus (**d**–**f**). Reprinted/adapted with permission from Ref. [184]. Copyright© 2022, Elsevier B.V.

#### 4.4.4. Rare Earth Metal Doping into ZnO

Doping rare earth metals into ZnO can improve the ability of the composite in trapping photogenerated carriers and reducing electron–hole recombination, which can enhance the photocatalytic antibacterial ability. In the rare earth doping process, f-orbital doping is the most common and efficient way, which can improve the photocatalytic activity by enhancing the adsorption of pollutants on the catalyst surface, while reducing the band gap energy to the visible light range [196]. Lanthanide ion doping is considered a versatile strategy to tune the optical response and improve the photocatalytic performance of ZnO. Lanthanides are composed of 17 elements in the periodic table, including Sc, Y, La, Ce, Pr, Nd, Pm, Sm, Eu, Gd, Tb, Dy, Ho, Er, Tm, Yb, and Lu. Lanthanides have attracted much attention due to their multifunctional properties resulting from their unique f-orbital structures, and due to the f-f or f-d intra-electron transitions, lanthanides are considered candidates for luminescent centers in doped materials, which is beneficial to prolonging the effective response time [197–199].

Doping ZnO with Ln3+ and Ce4+ ions can convert the magnetism from diamagnetism to ferromagnetism, improve the n-type conductivity, enhance the photo response, increase the concentration of free electrons in the CB, and increase the electron mobility [200–202]. A novel Z-type ZnO–CeO2-Yb2O3 heterojunction photocatalyst was prepared for the first time by Tauseef et al. [203], and its physical, photocatalytic, and antibacterial properties were investigated. Growth samples were tested for antimicrobial properties against *E*. *coli* and *S*. *aureus*. The effects of operating parameters such as catalyst dosage, dye concentration, and solution pH on the photocatalytic performance of the nanocomposites were investigated. The ZOIs of *S*. *aureus* and *E*. *coli* along with the standard antibiotic ciprofloxacin are shown in Figure 20a,b. The synthesized nanocomposites exhibited good activity against both bacteria with a ZOI > 6 mm, but higher activity against *E*. *coli* with a ZOI of 14 mm shown in Figure 21a,b. Positively charged heavy metal ions such as Zn2+, Ce4+, and Yb3+ can be released from the surface of the nanocomposite to interact with negatively charged microbial cell membranes. The entry of these metal ions into the cell membrane reduces the capacity and permeability of proteins, which in turn leads to the death of microorganisms such as bacteria and viruses. The above antibacterial action mechanism can be visualized in Figure 22C. In conclusion, the nanocomposites doped with rare earth ions are effective materials for preventing diseases caused by *S*. *aureus* and *E*. *coli*.

**Figure 20.** Antibacterial effect of ZnO-CeO2-Yb2O3 nanocomposites: (**a**) gram-positive *Staphylococcus aureus* bacteria, (**b**) gram-negative *Escherichia coli* bacteria at different concentrations, and (**c**,**d**) Standard antibiotics Ciprofloxacin against *S. aureus* and *E. coli*. Reprinted/adapted with permission from Ref. [203]. Copyright© 2022, Elsevier Masson SAS.

**Figure 21.** Comparison of the zone of inhibition (ZOI) against different species of bacteria (**a**,**b**), mechanism of antibacterial activity of ZnO–CeO2-Yb2O3 nanocomposite (**c**). Reprinted/adapted with permission from Ref. [203]. Copyright© 2022, Elsevier Masson SAS.

**Figure 22.** SEM images of various prepared CS/PVA/ZnO-related materials and precursors: (**A**) pure chitosan, (**B**) ZnO, and (**C**) CS/PVA/ZnO. Reprinted/adapted with permission from Ref. [204]. Copyright© 2022, Elsevier B.V.

#### 4.4.5. Organic Antimicrobial Agents Doping into ZnO

Studies have shown that the composites obtained by co-doping and fusion of organic antibacterial agents and ZnO nanoparticles exhibited stronger antibacterial activity than ZnO nanoparticles alone [205,206]. The organic antimicrobial agents are usually immobilized or embedded on the ZnO surface. Taking chitosan (CS) as an example, it is an abundant natural biopolymer derived from the deacetylation of chitin in crustacean shells and can be made into films, fibers, beads, and powders. Cationic polymers are generally antimicrobial [207,208]. In general, antibacterial activity depends on molecular weight (Mw), degree of deacetylation, temperature, and solution pH [209,210].

Gutha et al. [204] used CS and ZnO as raw materials to prepare a new composite material chitosan/poly(vinyl alcohol)/zinc oxide (CS/PVA/ZnO), which was used as a novel antibacterial agent with wound healing properties. CS/PVA/ZnO was proved to be an effective antibacterial nanomaterial after being analyzed by various characterization methods. SEM images of various prepared CS/PVA/ZnO-related materials and precursors are shown in Figure 22A–C. The surface of sole chitosan was obtained to be smooth. The sole ZnO nanoparticles showed nanosheet-like morphology. The surface of sole CS/PVA/ZnO microbeads presents a certain pore structure, and the surface of the microbeads is rough, which is conducive to exerting the ability of photocatalytic antibacterial. The antibacterial activities of CS, CS/PVA, and CS/PVA/ZnO are shown in Figure 23. The diameter of the inhibition zone against *E*. *coli* cultures (G-) was 10 mm in the CS group, 14 mm in the CS/PVA group, and 19 mm in the CS/PVA/ZnO group (Figure 23A). Likewise, *S*. *aureus* cultures (G+) had a diameter of 12 mm in the CS group, 15 mm in the CS/PVA group, and 20 mm in the CS/PVA/ZnO group (Figure 23A).

**Figure 23.** The antibacterial activities of CS, CS/PVA, and CS/PVA/ZnO: (**A**) Diameter of zone of inhibition on *E. coli* culture (gram negative bacteria); (**B**) Effect of mammalian cell viability of freshly prepared CS, CS/PVA, and CS/PVA/ZnO on MG63 (human osteosarcoma cell line) and (**C**) MCF7 (Human breast cancer cell line) cells. Reprinted/adapted with permission from Ref. [204]. Copyright© 2022, Elsevier B.V.

#### **5. Future Scope and Conclusions**

#### *5.1. Future Scope*

Although ZnO-based nanomaterials have been applied in organic pollutants removal from water and conduct antibacterial reactions in water, there is still plenty of space for improvement. The following points are the aspects that can be improved and strengthened in the application process of ZnO-based nanomaterials in the future:


#### *5.2. Conclusions*

This manuscript is based on recent developments in antibacterial water treatment with ZnO-based nanomaterials. Due to the increasing global requirements for water environment quality and drinking water quality, especially for the prevention and control of various epidemics, new ideas and directions are provided for our study. Therefore, the existence of bacteria and harmful microorganisms in water is introduced in detail, and various commonly used antibacterial agents and antibacterial methods are summarized. In conclusion, different morphologies of ZnO-based nanomaterials can be effectively used against various Gram-positive and Gram-negative strains by physicochemical interactions with bacterial cells. Cell membrane damage and biocidal activity are thought to be triggered by the collective action of chemical and physical interactions. Chemical interactions leading to the production of ROS and H2O2 and the release of Zn2+ ions from ZnO solubility have been proved to be the main cause of the above activities. Subsequently, based on this theory, an in-depth study of the antibacterial mechanism was carried out. Finally, the review also summarizes the following synthetic strategies to improve the antibacterial properties of ZnO: (1) doping of alkaline earth metals to ZnO; (2) doping of transition metals to ZnO; (3) doping of noble metals to ZnO; (4) doping of rare earth metal to ZnO; and (5) loading organic antimicrobial agents.

It can be expected that the antibacterial potential of ZnO-based nanomaterials in water treatment is very promising. Studies on ZnO-based nanomaterials continuously increased in recent years, although they still have many aspects that can be improved. In the future, we can expect more perfect ZnO-based nanomaterials to be prepared to solve more antimicrobial-related problems in water treatment.

**Author Contributions:** Z.X.: Investigation, Writing—Original Draft. Q.H.: Investigation. S.W.: Investigation. X.H.: Supervision, Writing—Reviewing and Editing. Z.F.: Supervision, Writing—Reviewing and Editing. X.X.: Supervision, Writing—Reviewing and Editing. X.Z.: Supervision, Writing—Reviewing and Editing. All authors have read and agreed to the published version of the manuscript.

**Funding:** This research was supported by the National Key Research and Development Plan, China (2019YFC1907204). We are grateful for the test services from the Analytical and Testing Center of Northeastern University.

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

**Informed Consent Statement:** Not applicable.

**Data Availability Statement:** Not applicable.

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

#### **References**


## *Review* **A Review of Treatment Techniques for Short-Chain Perfluoroalkyl Substances**

**Yang Liu 1,\*, Tingyu Li 1, Jia Bao 1,\*, Xiaomin Hu 2, Xin Zhao 2, Lixin Shao 1, Chenglong Li <sup>1</sup> and Mengyuan Lu <sup>1</sup>**


**Abstract:** In recent years, an increasing amount of short-chain perfluoroalkyl substance (PFAS) alternatives has been used in industrial and commercial products. However, short-chain PFASs remain persistent, potentially toxic, and extremely mobile, posing potential threats to human health because of their widespread pollution and accumulation in the water cycle. This study systematically summarized the removal effect, operation conditions, treating time, and removal mechanism of various low carbon treatment techniques for short-chain PFASs, involving adsorption, advanced oxidation, and other practices. By the comparison of applicability, pros, and cons, as well as bottlenecks and development trends, the most widely used and effective method was adsorption, which could eliminate short-chain PFASs with a broad range of concentrations and meet the low-carbon policy, although the adsorbent regeneration was undesirable. In addition, advanced oxidation techniques could degrade short-chain PFASs with low energy consumption but unsatisfied mineralization rates. Therefore, combined with the actual situation, it is urgent to enhance and upgrade the water treatment techniques to improve the treatment efficiency of short-chain PFASs, for providing a scientific basis for the effective treatment of PFASs pollution in water bodies globally.

**Keywords:** low carbon; short-chain PFASs; water treatment; adsorption; advanced oxidation

#### **1. Introduction**

Since the 1950s, perfluoroalkyl substances (PFASs) have been widely used in industrial production and commercial products, involving chrome plating, foam extinguishing agents, aviation hydraulic oil, and food packaging paper [1,2]. They are a class of manmade chemicals with all the hydrogen atoms on the carbon skeleton replaced by fluorine atoms, together with a terminal functional group [3]. Abbreviations for different PFASs are shown in Table 1. Due to the strong energy of the C-F bond (536 kJ/mol), PFASs possess exclusive physio-chemical characteristics, including environmental persistence, extraordinary resistance to both environmental and biological degradation, high thermal and chemical stability against oxidation, photolysis, and hydrolysis reactions, hydrophobicity and oleophobicity, as well as multiple toxicities [4]. Moreover, many PFAS (precursors) can easily degrade into persistent PFAS (acids). Therefore, long-chain PFASs (C8-C14) and their sodium, as well as ammonium, salts were added into the candidate list of regulatory substances in the EU, and PFOA and PFOS were added in the Stockholm Convention on Persistent Organic Pollutants (POPs) list [5]. With the ban of long-chain PFASs, shortchain PFASs (PFCAs < C8, PFSAs < C7) have been produced and used as substitutes in large quantities.

With the improvement of modern analytical techniques such as high-resolution mass spectrometry in non-target and suspect screening approaches in recent years, the researchers

**Citation:** Liu, Y.; Li, T.; Bao, J.; Hu, X.; Zhao, X.; Shao, L.; Li, C.; Lu, M. A Review of Treatment Techniques for Short-Chain Perfluoroalkyl Substances. *Appl. Sci.* **2022**, *12*, 1941. https://doi.org/10.3390/app12041941

Academic Editor: Dino Musmarra

Received: 18 January 2022 Accepted: 9 February 2022 Published: 12 February 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/).

found that long-chain PFASs tended to be adsorbed by solid matter (soil, sediment, etc.), which made them less mobile. However, short-chain PFASs showed a high polarity and solubility, which rendered them difficult to be removed through environmental adsorption and water treatment processes, thus contributing to long-term mobility in the water cycle through migration of surface water, groundwater, and natural and urban water systems [6–9]. However, short-chain PFASs show similar properties with long-chain congeners, including being persistent, bioaccumulative, and toxic to a certain extent [10]. Therefore, short-chain PFASs could be classified as persistent and mobile organic compounds (PMOCs) [6].

**Table 1.** Abbreviation for different PFASs.


Large quantity usage of short-chain PFASs could lead to accumulation in the water cycle and water pollution, thereafter threatening drinking water quality. For instance, the concentrations of PFBS and PFBA in Tangxun Lake of Wuhan in China reach up to 3.66 μg/L and 4.77 μg/L, respectively [11]. There was a serious PFBS contamination in groundwater around fluorochemical plants in Fuxin in China, with a concentration up to 31 μg/L [12], exceeding 10 times the health risk limits (HRLs) in drinking water (3 μg/L) issued by the Minnesota Department of Health (MDH). Thus, it is an urgent incident to effectively develop suitable water treatment techniques to regulate short-chain PFAS contaminations in waters.

At present, activated carbon (AC) adsorption and ozonation are the commonly applied techniques for organic contaminant elimination in drinking water treatment. However, AC presented low efficiency for very polar compounds [6]. Meanwhile, ozonation generally exhibited poor reactive activity to polar compounds containing acidic functional groups [13], which might be a source of smaller and more polar by-products than the parent compounds [14]. Consequently, the removal efficiency of conventional treatment for short-chain PFASs was undesirable. Moreover, various techniques presented different removal efficiencies for short-chain PFASs, even requiring high energy consumption, strict operating conditions, and releasing a large number of by-pollutants [15].

Recently, the C40 Cities Network of 91 large cities committed to low carbon infrastructure to ensure carbon emissions peak by 2020 and almost halve by 2030 [16–18]. In 2020, China announced that it would take more aggressive policy measures to achieve peak carbon dioxide emissions by 2030 and achieve carbon neutrality by 2060 [19]. Specifically, the regulation instruments include applying forced power to administer high energy consumption and emission of pollutants [20,21]. Therefore, it is significant to explore effective low carbon treatment techniques to eliminate short-chain PFAS contaminations under convenient conditions.

This review aims to select suitable treatment techniques for short-chain PFASs. To achieve this aim, the low carbon treatment techniques for PFASs involving adsorption, electrochemical oxidation, photocatalytic oxidation, membrane separation, pyrolysis, and ultrasonic chemical degradation, and their individual removal efficiency, operating conditions, and removal mechanisms were systematically summarized. Thereafter, based upon

the characteristics of short-chain PFASs, the suitable treatment techniques were determined by the comparison of the applicability, as well as the pros and cons of various techniques. This will provide a scientific basis for the effective treatment and regulation of short-chain PFAS contaminations in different waters.

#### **2. Methodology of Literature Sources**

To obtain an overview of short-chain PFAS chemical usage, the present review initially focused on the risk profiles and risk management assessments. Reports that addressed fluorosurfactants and fluoropolymers were also involved. Literature related to certain use categories was retrieved for more information either on the substances used, or to understand why PFAS are, or were, necessary for a given use.

In addition, databases, patents, information from PFAS manufacturers, and scientific studies were examined via "Web of Science", "PubMed", and "CNKI". The retrieved keywords involved but were not limited to "per- and polyfluoroalkyl substances", "PFAS", "PFAAs", "short-chain PFAS", "water treatment", "adsorption", "anion-exchange", "advanced oxidation", "persistent organic pollutant", "emerging contaminant", "low carbon". The searches were not exhaustive in any of the sources described, and there are still many more reports, scientific studies, patents, safety data sheets, and databases with information on the usage of PFASs than the ones cited here [22].

#### **3. Treatment Techniques for Short-Chain PFASs**

#### *3.1. Adsorption Technique*

Low carbon technique of adsorption could remove contaminations effectively in the waters, which were widely applied in the treatment of PFASs. This technique uses a porous solid as the adsorbent to adsorb one or several contaminations in the wastewater (WW) that does not change physicochemical property, thus achieving the purification purpose [4]. The commonly used adsorption materials mainly include carbons, anion-exchange resins, flocculants, etc. The key influencing factors of adsorption efficiency involved ionic strength, pH, organic matter (OM) concentration, physicochemical properties of PFASs in solution, and adsorbent characteristics (such as particle size, porosity rate, and functional group on surface). Comparisons on the adsorption capacities of different adsorbents for short-chain PFASs are shown in Table 2.

#### 3.1.1. Adsorption by Carbon-Based Adsorbents

AC adsorption. AC is widely used to remove contaminants in wastewater due to low cost, high efficiency, and convenient operation [4]. Ochoa-Herrera et al. adopted granular activated carbon (GAC) to adsorb PFBS, with an adsorption capacity of 98.7 mg/g [23]. Hansen et al. used GAC and powdered activated carbon (PAC) to carry out batch adsorption experiments for short-chain PFASs; PAC achieved higher removal rates within 10 min (20–40% for GAC, 60–90% for PAC) due to shorter internal diffusion distance and higher BET surface area (SBET) of PAC [24]. This phenomenon demonstrated that the particle size of AC was a significant factor affecting the removal efficiency of short-chain PFASs, and smaller particle sizes presented superior removal efficiency. The effect of pore size on the removal efficiency of bamboo-derived AC (BdAC) and coal-based AC (microporous type) for PFHxA were investigated, revealing that the BdAC adsorption capacity was 13-fold lower than microporous ACs [25,26]. In addition, the GAC was explored to eliminate different carbon chain length PFASs [27], the removal efficiency of short-chain PFCAs was lower than congeners PFSAs, in which the removal rates of PFBA, PFPeA, PFHxA, and PFHpA were all below 19%, especially 10% for PFBS. Furthermore, the regeneration capacity of adsorbed AC was also inferior when eluting. Therefore, it is necessary to explore new catalysts to modify AC for promoting adsorption and regeneration capacity [1].

Biocarbon adsorption. Biocarbon is a pollution-free solid biofuel produced by pyrolysis of biomass under aerobic or anaerobic conditions, which contains abundant voids, high carbon content, and high calorific value. The nature of biocarbon is different due to

various raw materials and parameters in the production process [28]. Inyang and Dickenson [29] explored the adsorption capacity of hardwood biochar (HWC) and pinewood biochar (PWC) for PFBA and PFOA, demonstrating that the PFBA adsorption capacity was 3–4 times lower than PFOA. Meanwhile, the removal efficiency of HWC vaporized at 900 ◦C could be improved due to high SBET. However, the batch adsorption kinetics experiments showed that the removal efficiency of GAC for PFBA was superior to HWC, indicating a low adsorption capacity of biocarbon for short-chain PFASs.

Carbon nanotube (CNTs) adsorption. CNTs have the advantages of easy reaction process control, convenient operation, and low cost of raw material. Deng et al. [30] used single-walled carbon nanotubes (SWCNT) and multi-walled carbon nanotubes (MWCNT) to remove short-chain PFASs (PFBA, PFHxA, PFHpA, PFBA, and PFBS) and long-chain PFASs (PFOA, PFHxS, and PFOS) under neutral conditions. It showed that 95% of PFOS and PFOA were removed by SWCNT within 5 h, but only 7.5% of PFBA was eliminated by SWCNT within 48 h, and PFSAs were more easily adsorbed than PFCAs. Moreover, in the comparison study of MWCNT functional groups, owing to the deprotonation that occurred on the carboxyl and hydroxyl functional groups, contributing to lower hydrophobicity and more negative surface potential, the adsorption efficiency of non-functional MWCNT for short-chain PFASs was improved. Therefore, based upon the above studies, the adsorption efficiency and technical maturity of CNTs were similar to ACs.

The main adsorption mechanism of carbon materials was hydrophobic and electrostatic effects, as well as possible hydrogen bonds and covalent bonds effects [4]. The hydrophobic effect would be improved with an increased C-F chain length; thus, longchain PFASs were better adsorbed than short-chain congeners. Meanwhile, the removal efficiency of adsorbents for PFASs was also depended on the terminal functional groups; for instance, the removal efficiency of PFSAs was better than PFCAs. In addition, the electrostatic attraction could occur between the anionic PFASs and positive charge adsorbents. Therefore, the changes of ionic strength involving cations or anions and pH in the solution would influence the adsorption efficiency. For example, the increase of ionic strength caused by monovalent or divalent cations (Na+, K+, Ca2+, etc.) might enhance PFAS removal efficiency, while the pH increase would reduce the adsorption capacity of most adsorbents [31,32]. However, the electrostatic repulsion between anionic PFASs and a negatively charged adsorbent could be overcome by the hydrophobic effect of the C-F chain [33]. Therefore, the removal of short-chain PFASs was mainly dependent on the electrostatic effect, while long-chain PFASs mainly tended to hydrophobic effects. Adsorption mechanisms of carbon-based adsorbents for PFASs removal are shown in Figure 1.

**Figure 1.** Adsorption mechanism of carbon-based adsorbents for PFAS removal.

#### 3.1.2. Anion-Exchange Resin Adsorption

Resin adsorption gradually attracted researchers' attention due to its strong adsorption, regeneration ability, and convenient operation. The carbon chain length (or hydrophobicity) and terminal functional groups of PFASs could influence the adsorption of anion-exchange resin for PFASs. Maimaiti et al. explored the adsorption of large pore anion-exchange resins about IRA910 for single PFASs (PFBA, PFHxA, PFOA, PFBS, PFHxS, and PFOS), showing that the chain length had a great influence on PFCAs adsorption compared with PFSAs [34]. The adsorption efficiency of PFSAs was better than PFCAs, in which the optimum adsorption efficiency of PFBS could be up to 1023.32 mg/g. Moreover, in order to investigate the treatment effect and regeneration capacity of anion-exchange resin, Du et al. eluted the IRA67 resin that saturated adsorption with PFOA and PFHxA using the mixture solution of NaCl and methanol, the recovery could achieve 98% and 40%, respectively. This phenomenon indicated that short-chain PFASs were difficult to remove from the resin. In addition, the properties of ion-exchange resin also played an effect on the adsorption for short-chain PFASs [25]. For instance, the adsorption capacity of the ion-exchange resin was superior to the non-ion exchange resin, and most of them were better than ACs.

The mechanism of anion-exchange resin mainly included hydrophobic effects, electrostatic effects, and ion-exchange effects. Generally, the resin with stronger hydrophobicity possessed a virtuous adsorption ability; however, the regeneration capacity was deprived by contrast [4]. Furthermore, the adsorption of short-chain PFASs might be influenced by pH via changing the resin surface potential or morphology [34]. For instance, in various ranges of pH, strong base anion (SBA) resin was impregnable due to its ionization form. In contrast, weak base anion (WBA) resin was influenced significantly [31], which could take effect when the amine group was protonated under acidic conditions. Moreover, the main mechanism for short-chain PFAS removal was single-molecule anion-exchange by the analysis of transmission electron microscope; thus, the inorganic anions in solution would compete with PFAS anions for the ion-exchange sites and then decrease the removal efficiency. The ion-exchange reaction equations are shown in (1) and (2), where [R4N+] and [R3N] indicate the ion-exchange site [15].

$$\text{[[R}\_4\text{N}^+]\text{Cl}^- + \text{PFAS}^- \quad \leftrightarrow \quad \text{[R}\_4\text{N}+\text{JPFAS}^- + \text{Cl}^-\tag{1}$$

$$\text{[[R}\_3\text{N]} + \text{H}^+ + \text{PFAS}^- \quad \leftrightarrow \quad \text{[[R}\_3\text{NH}^+\text{]}\text{PFAS}^-\tag{2}$$

#### 3.1.3. Coagulation and Electrocoagulation

Coagulation possessed the advantage of low price and high adsorption efficiency [4]. Deng et al. [35] discovered that the removal efficiency of PFOA exceeded 90% when the dosage of polymer aluminum chloride (PACl) was 10 mg/L. However, the removal efficiency of PFOA was reduced significantly when multiple PFASs existed simultaneously, and the removal rates followed the order of PFBA > PFHxA > PFOA > PFDoA > PFOS, which demonstrated that the removal efficiency of short-chain PFASs was superior, compared with long-chain congeners.

Electrocoagulation received widespread attention because of its higher removal efficiency for PFASs and short treatment period [4]. The electrocoagulation technique mainly produced a large number of cations and then generated flocs by sacrificing the anode; thus, the dissolved contaminants could be purified by condensation and adsorption of flocs, which was subsequently carried to the surface of the solution by the H2 and O2 produced by the electrodes through electrical floating. The Electrocoagulation mechanism for PFAS removal is shown in Figure 2. Liu et al. [36] adopted the periodically reversing electrocoagulation (PREC) technique to treat contaminated groundwater around fluorochemical plants, indicating that the PREC was effective for the removal of PFASs with different lengths of carbon chains. Subsequently, the above group approved that the PREC technique with Al-Zn electrodes for multiple PFSA removal was impactful, the removal rates of PFBS, PFHxS, and PFOS could reach up to 87.4%, 95.6%, and 100% within 10 min under

optimal conditions (12.0 V, pH = 7, 400 r/min) [37]. In general, the removal mechanism of electrocoagulation was a hydrophobic effect and might exist with electrophoretic, polarized, and electric fields [38]. In addition, several influencing factors involving current density, inorganic ions, and OM were also crucial for the removal of short-chain PFASs.

**Figure 2.** Electrocoagulation Mechanism for PFASs removal.



#### 3.1.4. Adsorption with Other Materials

In recent years, the performance and construction of new materials that could be controlled and modified easily were synthesized by researchers. Wang et al. used a covalent triazine-based framework (CTF) to eliminate PFBS, with an adsorption capacity of 92.03 mg/g [25]. Subsequently, Zaggia et al. explored the adsorption capacity of AC, CTF, and IRA910 for PFBA and found that the order followed the rules of micropore AC < CTF < IRA910, which demonstrated the adsorption mechanism was an electrostatic effect between the triazine group and the PFAS anion head [32]. Ionic strength might be another major factor of other materials that influenced the adsorption efficiency. For example, the adsorption efficiency of poly-styrene carboxylic acid (PS-COOH) for shortchain PFCAs in seawater was better than river water due to the large ionic strength in seawater [39].

Therefore, in various kinds of adsorption materials, the adsorption efficiency of anion exchange resin and electrocoagulation for short-chain PFASs were remarkable. However, the elution ability of the anion exchange resin was inferior when the short-chain PFASs were adsorbed, and the flocs of electrocoagulation were still required for further treatment. Secondly, AC owned the property of low cost and high removal efficiency, while the regeneration ability was unsatisfactory. Finally, new adsorption materials could improve electrostatic attraction effectively, but the application should be further explored.

#### *3.2. Advanced Oxidation/Reduction Techniques*

Advanced oxidation/reduction techniques have been used for the degradation of PFASs, with the advantages of high conversion efficiency and simple operation, and some techniques could achieve complete mineralization. However, these techniques generally put emphasis on long-chain PFASs, including PFOA and PFOS. Whether these techniques could remove short-chain PFASs was still lacking in studies. This section provides a systematic summary of the application of the degradation of short-chain PFASs about the techniques of electrochemical oxidation and photocatalytic degradation, etc.

#### 3.2.1. Electrochemical Oxidation

Electrochemical oxidation is an emerging advanced oxidation technique due to the advantages of high removal efficiency, strong oxidative ability, and low energy consumption. This technique was found to degrade long-chain PFASs effectively. The most commonly used oxidation anodes included Ti/SnO2, Ce/PbO2, boron-doped diamond (BDD) electrodes, and their modified electrodes. Comparisons on the removal efficiency of short-chain PFASs by different electrode materials are shown in Table 3.


**Table 3.** Removal efficiency of short-chain PFASs by different electrode materials.

Niu et al.'s research group found that the degradation efficiency of Ti/SnO2-Sb for 100 mg/L PFOA could reach 98.8% when the current density was 40 mA/cm2, pH at 5, and the electrolyte with 10 mmol/L NaClO4 [43]. Subsequently, the research group explored the Ti/SnO2-Sb/PbO2-Ce electrode to degrade short-chain PFASs of PFHpA and PFBA, with removal rates of 97.9% and 31.8%, respectively. This might be related to the high resistance of short-chain PFASs and the co-existence of multiple PFASs [40]. In addition, Soriano et al. used a BDD electrode to remove high concentrations of 870 mg/L PFHxA in a solution containing OM and inorganic salt, finding that the removal rate was 98% within 2 h when the current density was at 100 mA/cm2, and energy consumption was 45 Wh/L [41]. Liao et al. explored a Si/BDD electrode for the elimination of high concentration PFBS (>150 mg/L) at low current density; the removal rate was 90% within 1 h [42]. Based upon the above studies, the BDD and its modified electrodes achieved remarkable results compared with other electrodes involving SnO2 and PbO2. However, the base materials of Si were too weak and performed unfortunate electrical conductivity. Therefore, high manufacturing costs and lack of suitable base materials limited the large-scale application of BDD electrodes. At present, it is important to discover a cheap and stable base material for the industrial application of BDD electrodes [37].

The degradation of PFASs on the electrode surface was related to electron transfer and free radical oxidation; the proposed pathways for electrochemical oxidation of PFOS in water are shown in Figure 3.

**Figure 3.** Proposed pathways for electrochemical oxidation of PFOS in water.

Initially, the electrons are transferred from the terminal functional group of PFASs to the anode driven by the electric field and formed PFAS radicals (·CnF2n+1COO or ·CnF2n+1SO3), the extremely unstable PFAS radicals undergo decarboxylation or desulfation to form ·CnF2n+1. The generated ·CnF2n+1 could react with H2O, OH−, or ·OH, and finally -CF2 groups were cut down gradually to generate short-chain PFASs with the specific reaction equations shown in (3)–(7) [44,45].

$$\text{C}\_{\text{n}}\text{F}\_{\text{2n}+1}\text{COO}^{-}\rightarrow \begin{array}{rcl} \cdot \text{C}\_{\text{n}}\text{F}\_{\text{2n}+1}\text{COO} + \text{e}^{-} \end{array} \tag{3}$$

$$\text{C}\_{\text{n}}\text{F}\_{\text{2n}+1}\text{SO}\_{3}^{-} \quad \rightarrow \quad \cdot \text{C}\_{\text{n}}\text{F}\_{\text{2n}+1}\text{SO}\_{3} + \text{e}^{-} \tag{4}$$

$$\text{Cr}\_{\text{n}}\text{F}\_{\text{2n}+1} + \text{H}\_{2}\text{O} \quad \rightarrow \quad \cdot \text{CrF}\_{\text{2n}+1}\text{OH} + \text{H}^{+} \tag{5}$$

$$\rm C\_nF\_{2n+1}OH \quad \rightarrow \quad C\_{n-1}F\_{2n-1}COF + H^+ + F^- \tag{6}$$

$$\text{C}\_{\text{n}-1}\text{F}\_{\text{2n}-1}\text{COF} + \text{H}\_{2}\text{O} \quad \rightarrow \quad \text{C}\_{\text{n}-1}\text{F}\_{\text{2n}-1}\text{COOH} + \text{H}^{+} + \text{F}^{-} \tag{7}$$

Plenty of studies have demonstrated that the degradation efficiency of short-chain PFASs was still lower than long-chain congeners. In order to improve the removal efficiency of short-chain PFASs and reduce the energy consumption effectively, the studies of influence factors, including the current density, pH, electrolyte, OM, and microorganisms in solution, must be carried out. Meanwhile, the electrode modification could improve the removal efficiency and the defluoridation efficiency of PFASs, as well as improve the electrode life successfully, but the metal doping on electrodes might lead to partial contamination during treatment processes [4].

#### 3.2.2. Photocatalytic Degradation

Direct photolysis. Photocatalytic degradation is an advanced oxidation technique that can mineralize target compounds by UV light and photocatalysts. However, PFASs cannot absorb light over 220 nm directly, and the degradation rate of PFBA and PFPeA was only 16.3% and 24.3% by direct photolysis, respectively [15]. Hori et al. found that PFPeA could strongly absorb the light from the vacuum UV region to 220 nm due to the degradation of PFASs under VUV irradiation (<190 nm) by the active substances produced by homolysis and ionization of water, involving a hydrated electron (eaq−), hydrogen radical (H·), and hydroxyl radical (·OH) [46]. Therefore, how to use catalysts to promote the generation of active substances about free radicals for the photodegradation of PFASs was imperative. The photocatalytic efficiency of PFASs with different catalysts is shown in Table 4.

The sulfate radical (SO4 •−) produced from persulfate (S2O8 <sup>2</sup>−) under UV irradiation was observed to be more effective for PFASs degradation over ·OH. Therefore, S2O8 <sup>2</sup><sup>−</sup> was used as an oxidant to degrade PFAS frequently [47]. For instance, Hori et al. found that the short-chain PFCAs in aqueous solution were oxidized to CO2 and F− by SO4 •− when the concentration of S2O8 <sup>2</sup><sup>−</sup> was 50 mmol/L [46]. Subsequently, this group also found that Fe3+ was an effective catalyst for photodegradation, and the photodegradation efficiency of PFBA and PFPeA were improved to 49.9% and 64.5% due to strong light absorption from the complexation of PFASs with Fe3+ [47]. Water-soluble polyacid photocatalysts could also promote C-F bond cleavage of PFPrA, thereafter generated to TFA, CO2, and F−, but the energy consumption of the process was exorbitant [48].

Zero-valent iron (ZVI) reduction. Reductants including zero-valent iron particles or iodine salt could also enhance the photodegradation of PFASs. ZVI was used as a reductive agent for the photodegradation of PFASs due to its high reduction potential (−0.447 V) and reactive surface area [15]. For instance, Hori et al. observed that the degradation efficiency of short-chain PFSAs (C2-C6) could be up to 95% using ZVI, owing to the generation of iron oxide in the ZVI surface, which could interact with PFASs ions synergistically, thus promoting defluoridation efficiently [49].

Photocatalytic degradation. Photocatalysts were scarcely implemented in short-chain PFASs; the photodegradation efficiency could be deduced by examining the degradation intermediates and degradation efficiency of long-chain PFASs. For example, Panchangam et al. adopted TiO2 photocatalysts for 120 mmol/L of PFOA degradation under UV irradiation at 254 nm within 7 h, 97% of PFOA were converted to short-chain intermediates, including PFPrA, PFBA, PFPeA, PFHxA, and PFHpA, in which PFHpA reached the maximum at 5 h [50]. Currently, the widely used method for modifying TiO2 was doping with precious metals (Pt, Pd, Au) or other metals (Pb, Cu, Fe). Li et al. explored TiO2 doped with Pt to completely decompose 144.9 mmol/L PFOA within 7 h under UV irradiation at 365 nm; the defluoridation efficiency was 34.8%. It was 12.5-fold faster than unmodified TiO2 due to the deposited Pt particles could store excessive electrons and promote the electrons transferred to PFAS availably [51]. In addition, the composite materials were used to improve the photocatalytic degradation of PFASs, such as TiO2-MWCNT and TiO2-rGO. It was found that the degradation efficiency of PFOA could reach 100% when using TiO2-MWCNT after 8 h under irradiation of UV at 265 nm [52].

In addition to the above n-type photocatalysts of TiO2, p-type photocatalysts involving In2O3 and Ga2O3 have attracted widespread attention because they could enhance the degradation capacity of PFASs. However, p-type photocatalysts showed a strong dependence on the material shape and microstructure [15]. For example, the In2O3 porous microsphere had the highest photocatalytic activity on the degradation of PFOA, which was 74.7 times faster than TiO2. Similarly, short-chain intermediate PFHpA could be completely degraded within 3 h by the In2O3 nanoplates, whereas In2O3 nanocubes were much less effective. However, in the degradation process, the short-chain intermediates were still generated from PFOA degradation, and the concentration of intermediates was positively related to their carbon chain length [50,53].

New photocatalysts could decompose PFAS, but the information about the application for short-chain PFASs was still limited. According to the degradation data of long-chain PFASs, the degradation conclusions could be deduced as the following: photocatalytic degradation of PFASs was a gradual splitting decomposition of CF2. The reaction of breaking the chain generated short-chain intermediates. The short-chain products showed strong resistance to photocatalytic degradation [15].

During the photodegradation process, the catalyst dosage and pH were imperative influencing factors. In the low concentration range between 20–100 mmol/L of persulfate, the photochemical reactivity could be improved with the concentration increase. Whereas further increasing persulfate concentrations could result in saturation of the reaction rate, Because SO4 •− could react with persulfate or itself, this side reaction would reduce the degradation efficiency [15]. Moreover, a high concentration of hydrated hydrogen ions (H3O+) would quench eaq<sup>−</sup> under acidic conditions, thus contributing to the reactivity decrease due to the quantum yield of eaq− declining sharply. However, under alkaline conditions, the reactivity would be enhanced due to the quantum yield of eaq− increasing by the reaction of H· and OH<sup>−</sup> [54].


**Table 4.** Photocatalytic efficiency of PFASs with different catalysts.

#### *3.3. Other Techniques*

#### 3.3.1. Plasma Technique

Plasma is the collection of positive and negative electric particles that consist of electrons, ions, radicals, and neutral particles, which are electric and electroneutral, thus called the fourth state beyond gas, liquid, and solid electrically. Diverse from most AOPs and conventional techniques, plasma techniques could convert water into highly active substances, involving ·OH, O, H·, HO2·, O2 •−, H2, O2, H2O2, and eaq− [56]. When using the plasma technique for PFASs elimination, the degradation process was gradually reduced to intermediates, and then intermediates, perfluoroalkyl radicals, and perfluoro alcohols/ketones-perfluoroalkyl were oxidized [57]. However, this technique was hardly

applied and is invalid in the degradation for short-chain PFASs. For example, Takaki et al. adopted BaTiO3 iron beads as a filling medium to degrade C2F6; the degradation efficiency was only 20% [58]. Moreover, short-chain PFASs, fluorine ions, and CO2 by-products were produced during the plasma treatment of PFASs.

#### 3.3.2. Thermolytic and Sonochemical Degradation

In the recent decade, thermolytic and sonochemical degradation have attracted wide attention. Tsang et al. [59] found that 99% of CF4 could be removed under urban incineration conditions (about 850 ◦C) and assumed the pyrolysis of 800–900 ◦C could resolve long and short-chain PFASs efficiently. Krusic et al. [60] explored the gas phase decomposition of PFOA in quartz tubes at 355–385 ◦C, finding that PFOA retained thermal stability below 300 ◦C but completely degraded at 370 ◦C after 360 min. However, the by-products of small molecule substances were generated continuously, which demonstrated the greater resistance of short-chain PFASs. Subsequently, Campbell et al. explored pre-concentrating PFHxA on GAC, which could improve the thermal mineralization rate from 46% to 74% [61]. Furthermore, in order to investigate the effect of ultrasonic chemistry on the degradation of PFASs with different carbon chain lengths, Campbell et al. investigated the degradation efficiency on six kinds of PFASs at 358 kHz, the order was PFOA > PFHxA > PFBA and PFOS ≈ PFHxS > PFBS [62]. Similarly, Fernandez et al. found that the degradation efficiency of ultrasonic chemicals raised with carbon chain length increased [63]. For one thing, long-chain PFASs with strong hydrophobicity were inclined to be adsorbed on the gas-liquid interface for thermal decomposition or oxidation degradation. For another, short-chain PFASs were more difficult to be defluorinated than long-chain congeners. Generally, these techniques could mineralize short-chain PFASs, but nongreen environmental factors about the generous discharge of CO2 and high energy consumption limited the development in actual application.

#### 3.3.3. Membrane Separation

In recent years, low carbon treatment of membrane separation techniques, including nanofiltration (NF) and reverse osmosis (RO), have made great advances. During the rejection in membrane processes, the molecular size and structure of PFASs were considered key factors. For instance, NF membranes could reject more than 96% of PFHxA (μg/L to mg/L) under neutral pH, whereas the rejection rate of shorter-chains PFASs about PFBS decreased to 69% attributable to the small molecular size [64]. Furthermore, charge, hydrophobicity, pH, and dipole moment might also affect the solute-membrane interactions and thus the rejection efficiency of PFAS. It was shown that pH reduction could increase the membrane rejection efficiency of PFHxA. Moreover, the presence of the ions generally suppressed the electrical repulsion, but the exclusion efficiency was enhanced with increasing ionic strength, which indicated that the exclusion of membrane size was the dominant factor. In the practical application, the rejection efficiency of short-chain PFASs would be easily influenced because of membrane contamination. In addition, membrane separation techniques could produce a high concentration of PFASs and still require subsequent treatment or disposal.

#### 3.3.4. Bioremediation Techniques

Kwon et al. degraded PFOS (1400–1800 μg/L) by cultivating *P. aeruginosa* microorganism with the removal rate of 67% [65]. However, short-chain PFASs were difficult to degrade by common microbes. For example, the concentration of PFBS in the effluent of sewage treatment plants remained unchanged or increased after conventional activated sludge or biofilm bioreactors. That is, short-chain PFASs showed high resistance to various activated sludge systems [5]. While in plant tissues, both short and long-chain PFASs all have a tendency to accumulate. Recent studies have shown that the biological accumulation of PFASs followed a U-type trend. The lowest hydrophobicity (e.g., PFBA and PFPeA) and the maximum hydrophobic species (e.g., PFNA and PFDA) presented the greatest absorb

efficiency [66]. Another study on the uptake and distribution of PFASs in maize showed that plant adsorption and distribution of PFASs were dependent on chain length, functional groups, and plant tissue. Generally, short-chain PFASs were transferred to the overground portion of plants, while long-chain PFASs were mainly transported to the root [67].

#### **4. Comparisons on Different Treatment Techniques**

So far, low carbon treatment techniques of short-chain PFASs have included adsorption, membrane separation, bioremediation, as well as degradation techniques relating to advanced oxidation, plasma, thermolytic, and sonochemical degradation. The above technologies could remove short-chain PFASs to a certain extent, but their treatment effects, operating conditions, removal mechanism, and applicability were quite different. The comparisons on treatment techniques for short-chain PFASs are presented in Table 5.

Adsorption was the utmost widely applied technique for short-chain PFASs, and its energy consumption could be nearly ignored besides the low energy cost of the adsorbent regeneration. The removal mechanisms of short-chain PFASs were mainly electrostatic action, hydrophobic effect, and ion exchange [4]. As shown in Table 5, this technique has the advantages of convenient operation, low carbon, low cost, and low energy consumption, as well as application with a wide concentration range of short-chain PFASs, especially trace levels. However, the technique has the drawbacks of a long adsorption period and low regeneration efficiency [27,36].

Advanced oxidation techniques, involving electrochemical oxidation and photocatalytic degradation, degraded short-chain PFASs primarily relying on active free radicals, which possessed the advantages of a short treatment period and low energy consumption [5,15]. However, the techniques were not suitable for trace levels of short-chain PFASs, and its low mineralization rate and subsequent CO2 generation rate from mineralization were also the main problems for achieving the low-carbon goals [43]. In addition, electrochemical oxidation could produce high expenses of electrode materials and the risk of electrode contamination. The photocatalytic degradation technique generated the problem of secondary pollution by catalyst addition and might thus be limited in actual applications [15].

The plasma technique was applied to the degradation of long-chain PFASs effectively, but few studies focused on its degradation of short-chain PFASs, and the energy consumption was still higher for complete mineralization [58]. The thermolytic and sonochemical techniques could achieve complete mineralization, but the energy consumption was too high, and the operating conditions were stringent [61]. The membrane separation technique could reject short-chain PFAS pollutants effectively, but membrane pollution and membrane flux instability were the main problems [64]. Bioremediation techniques could take advantage of their low carbon and environmentally-friendly processes, but they generated the problems of long remediation period and low efficiency, as well as inefficient short-chain PFAS elimination and remaining in organisms [65].

Based upon the analysis, the most widely used and effective method could be adsorption, followed by advanced oxidation. However, there were still limitations of removal efficiency in the application of eliminating short-chain PFASs. Since short-chain PFASs were more resistant to be adsorbed and degraded than long-chain congeners, PFASs of C1–C3 were barely degradable. Combined techniques might be developed based on concentration/recycling-degradation of short-chain PFASs in water bodies rather than degradation. The combination of adsorption and gas–liquid series electrical discharge treatment has been applied in the degradation of dyestuff, which achieved excellent removal efficiency compared with the adsorption alone [68].


#### **5. Conclusions and Future Research Recommended**

This paper showed that the adsorption, electrochemical oxidation, and photocatalytic degradation have certain removal effects on short-chain PFASs by comparisons on various treatment techniques. Considering the removal efficiency, treatment time, energy consumption, and cost, adsorption was the most widely applied technique for the effective removal of short-chain PFASs, which was suitable for a wide concentration range of pollution and to meet the low-carbon policy. Whereas, long adsorption period and unsatisfied regeneration ability were the main problems. The advanced oxidation techniques of electrochemical and photocatalytic activity could degrade short-chain PFASs, but low mineralization efficiency contributed to intermediates of short-chain PFASs, as well as abundant organic matter and CO2 and they were especially inappropriate to eliminate trace short-chain PFASs. Therefore, it was desirable to choose suitable techniques according to PFAS properties, as well as the advantages and disadvantages of various techniques.

The contaminations of short-chain PFASs have attracted much attention at present, while most studies still focused on the laboratory-scale treatment of long-chain PFASs, including PFOA and PFOS, and data on the treatment of short-chain PFASs is still absent. Therefore, future studies need to focus on the following topics: (1) Targeting long duration period and poor material regeneration ability for the adsorption technique; the functional groups of adsorbent materials need to be modified for improving electrostatic attraction and hydrophobic effects and enhancing the adsorption efficiency and elution of short-chain PFASs. (2) Innovative design for short-chain PFAS treatment by electrode and catalyst modification to develop advanced oxidation techniques, with high degradation efficiency in low-carbon and low-energy consumption, as well as adaptability for low-concentration PFASs. (3) With the increasing pollution of short-chain PFASs in drinking water, extensive emphasis should be placed on the development of advanced treatment techniques for actual groundwater and surface water, along with exploration of new adsorption materials, electrode materials, and catalysts, which could remove low concentrations of shortchain PFAS under the background of multiple substances co-existing in the actual waters. (4) Short-chain PFASs were more resistant to degradation than long-chain congeners, and PFASs of C1-C3 were barely degradable. Therefore, concentration/recycling of short-chain PFASs in water bodies should be considered rather than degradation. In addition, combined techniques might be developed based on concentration/recycling-degradation, such as adsorption and advanced oxidation, for the efficient removal of short-chain PFASs from actual waters.

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

**Funding:** This research was funded by the National Natural Science Foundation of China (No. 21976124 and No. 21507092), the Natural Science Foundation of Liaoning Province of China (No. 2019-ZD-0217), and the Liaoning Revitalization Talents Program (No. XLYC2007195).

**Acknowledgments:** Thanks for the financial support from the National Natural Science Foundation of China (No. 21976124 and No. 21507092), the Natural Science Foundation of Liaoning Province of China (No. 2019-ZD-0217), and the Liaoning Revitalization Talents Program (No. XLYC2007195).

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

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