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Article

High Salinity Tolerance of Zn-Rich g-C3N4 in the Photocatalytic Treatment of Chlorophenol Wastewater

by
Hongyu Chen
1,
Ying Wang
2,
Suiyi Zhu
3,
Xiaoshu Wang
4,
Jiancong Liu
4,
Lei Wang
4,5,*,
Wei Fan
1,* and
Yang Yu
6
1
School of Environment, Northeast Normal University, Changchun 130117, China
2
State Key Joint Laboratory of Environment Simulation and Pollution Control, School of Environment, Tsinghua University, Beijing 100084, China
3
College of Resources and Environment, Zhongkai University of Agriculture and Engineering, Guangzhou 510408, China
4
School of Materials and Environmental Engineering, Shenzhen Polytechnic University, Shenzhen 518055, China
5
Eco-Environmental Science Center (Guangdong, Hong-Kong, Macau), Guangzhou 510555, China
6
Biology Department, Hope College, 35 East 12th Street, Holland, MI 49423, USA
*
Authors to whom correspondence should be addressed.
Water 2024, 16(19), 2756; https://doi.org/10.3390/w16192756
Submission received: 26 August 2024 / Revised: 25 September 2024 / Accepted: 26 September 2024 / Published: 27 September 2024
(This article belongs to the Special Issue Sustainable Water Treatment and Contaminants Control: Technologies and Strategies)
Browse Figures
Versions Notes

Abstract

:
Organic saline wastewater has become a concern in recent decades due to its resistance to biological treatment and potential harm to municipal wastewater treatment plants. While photocatalytic methods have been used for treatment, they often lead to catalyst deterioration. The use of salt-tolerant catalysts presents a viable solution for treating organic saline wastewater. In this study, a Zn-rich g-C3N4 was synthesized, demonstrating excellent performance in removing 2,4-DCP and its derivatives from saline wastewater. More than 75.6% of 2,4-DCP was effectively removed with the addition of Zn-rich g-C3N4, nearly doubling the removal rate compared to pure g-C3N4 and those doped with Co, Ag, Mo, and Bi. Notably, the removal efficiency of 2,4-DCP slightly increased as salinity rose from 0.1 to 2.3 wt.%. Adding 0.1 g L−1 of Zn-rich g-C3N4 resulted in the removal of 2,4-DCP, 2-chlorohydroquinone, chloroacetophenone, and 2-chloropropionic acid by 99.3%, 99.8%, 98.2%, and 99.9%, respectively, from a real saline wastewater sample with 2.2 wt.% salinity, corresponding to a 67.7% removal of TOC. The EPR results indicated that Zn-rich g-C3N4 generated more free radicals compared to pure g-C3N4, such as·OH and Cl, to degrade organic contaminants. The degradation pathway revealed that 2,4-DCP was first dechlorinated into p-phenol and catechol, which were subsequently degraded into maleic acid/fumaric acid, trihydroxyethylene, acetic acid, oxalic acid, and other products. Furthermore, Zn-rich g-C3N4 demonstrated excellent stability and holds promising potential for applications in saline wastewater treatment.

1. Introduction

With the gradual expansion of the paper, printing, dyeing, pharmaceuticals, and pickled food industries, large volumes of complex, high-salinity organic wastewater have been generated [1,2]. Some of the wastewater usually contains high concentrations of toxic compounds, which greatly impact local environments. For example, it has been reported that wastewater from the production of the herbicide 2,4-dichlorophenoxyacetic acid contains 2,4-dichlorophenol (2,4-DCP) at concentrations ranging from 800 to 1200 mg L−1, while NaCl concentrations reach 5.5 to 6.5 wt.% [3]. Due to its high toxicity, ecological persistence, accumulation, and mobility, 2,4-DCP poses significant risks to ecosystems and public health, leading it to be classified as a priority pollutant by many countries [3]. However, the high salinity of this wastewater negatively affects both aerobic and anaerobic bacteria, therefore limiting the effectiveness of the biological treatment process [4].
Currently, several physical-chemical treatment processes, including the Fenton reaction [5], membrane filtration, evaporative crystallization [6], electrocatalysis [7], and photocatalysis, are considered advantageous for managing this type of wastewater. The removal of these recalcitrant organic contaminants without the negative effects of salinity is highly desirable, but materials with high surface energy tend to induce salt precipitation and crystallization, complicating processes for membrane filtration, electrocatalysis, and other processes [8,9,10]. It has been demonstrated that the decomposition of H2O2 catalyzed by Fe2+ produced ·OH (E0 = 2.70 VNHE) with strong oxidizing capabilities to non-selectively attack organic matter, and the dissolved oxygen gained electrons under catalytic conditions to produce ·O2− (E0 = 1.06 VNHE) to selectively actions against electron-rich groups [11,12]. Evaporative crystallization effectively facilitated the phase transfer of organic matter. Nevertheless, the Fenton reaction and evaporation crystallization required substantial reagent inputs and additional energy. Compared with the above, photocatalytic technology could continuously and efficiently decompose organic matter in saline wastewater while requiring lower energy consumption.
Various natural and synthetic photocatalysts have been developed to remove organic pollutants from wastewater. Among these, carbon nitride has emerged as a notable polymer material, featuring a heptazine ring structure and gaining attention in semiconductor applications over recent decades. Known for its stability and non-toxicity, carbon nitride is frequently used as a photocatalyst. Upon light irradiation, electrons in its highest occupied molecular orbital, composed of Npz orbitals, transition to the lowest unoccupied molecular orbital, consisting of Cpz orbitals. Compared to conventional materials such as TiO2 and SnO2 [13,14,15], the electron transition in carbon nitride requires less energy. As a result, carbon nitride is commonly employed to degrade a range of organic contaminants, including methylene blue, aniline, and tetracycline [16].
Furthermore, carbon nitride can be modified to fine-tune properties such as hydrophilicity, morphology, surface functional sites, and structural defects [17,18], which generally improve its catalytic performance in contaminant degradation [19]. However, wastewater is a complex, multi-component system, and salts present in the water often occupy the active surface sites of carbon nitride, inhibiting its adsorption and catalytic degradation of contaminants [20]. Unfortunately, the salt tolerance of g-C3N4 has been scarcely reported.
This study aimed to enhance the photocatalytic performance and salt tolerance of conventional carbon nitride catalysts through metal doping. By incorporating zinc, a Zn-rich g-C3N4 catalyst was developed, which demonstrated improved performance in treating saline wastewater. The effects of salinity and pH on the stability of the Zn-rich g-C3N4 were also evaluated.

2. Materials and Method

2.1. Chemicals and Materials

The wastewater was sampled at the discharge outlet of a pesticide company in Guangzhou, China. The ultrapure water required for the experiment with a resistivity of 14.60 MΩ∙cm. Zinc chloride (ZnCl2, 99%), melamine (C3H6N6, 99%), Sodium chloride (NaCl, 99.5%), Sodium hydroxide (NaOH, 97%), ethanol (C2H6O, 99.7%), 2,4-Dichlorophenol (2,4-DCP, 99.5%) and Methanol (CH4O, 99.9%) were supplied by Macklin Biochemical Co., Ltd. (Shanghai, China).

2.2. Catalyst Preparation

The catalyst was synthesized following the method of Oh et al. [18] with some modifications. First, 6.82 g of ZnCl2 and 25.20 g of melamine were mixed in a molar ratio of 4:1 in 300 mL of 0.35 wt.% HCl solution at 90–95 °C and stirred for 30 min. Next, the mixture was freeze-dried for two days, followed by calcination at 400–500 °C for 4 h to obtain a powdered material. Finally, the material was washed several times with water and ethanol and then dried overnight. The resulting pink powder was designated as Zn-rich g-C3N4 (Zn-C3N4). See Figure 1.

2.3. Photocatalytic Experiment

The prepared Zn-C3N4 catalyst was applied for the removal of 2,4-DCP from saline wastewater. First, 20 mg of g-C3N4 was mixed into 200 mL of wastewater containing 100 mg L−1 2,4-DCP and 0.1 wt.% salinity. The mixture was agitated at 500 rpm for 30 min in the dark while preheating the high-pressure mercury lamp (UVA365, 300 W, irradiance of 0.5 kW m−2) for 20 min. Next, the photocatalytic experiment was initiated by connecting the reactor, condensation, and mercury lamp units, and 1 mL wastewater samples were taken at regular intervals, followed by quenching with methanol. The residual 2,4-DCP concentration was analyzed using HPLC (LC-40, Shimadzu, Kyoto, Japan). Control experiments were conducted under varying conditions, including salt concentration (0.1–2.3 wt.%), pH (3–11), and 2,4-DCP concentration (20–200 mg L−1). The experiment was repeated three times, and the average results were reported.

2.4. Characterization

The prepared samples were characterized using several techniques. X-ray diffraction (XRD) was performed with a Shimadzu XRD-7000S (Japan), utilizing Cu-Kα radiation at a scanning rate of 10 °C min−1. Scanning electron microscopy (SEM, SU8000, Hitachi, Tokyo, Japan) was conducted with an accelerating voltage of 3–20 kV. Fourier transform infrared spectroscopy (FTIR) analysis was carried out with a Spectrum 3TM (PerkinElmer, Waltham, MA, USA), while X-ray photoelectron spectroscopy (XPS, ESCALAB 250Xi, Thermo Fisher, Waltham, MA, USA) used Al-Kα radiation (1486.6 eV) as the X-ray source. UV-visible diffuse reflectance spectroscopy (UV-Vis DRS, UV-2550, Shimadzu, Kyoto, Japan) was performed with BaSO4 as a reference material. Chloride release in solution was determined via ion chromatography (IC, 940, Metrohm, Riverview, FL, USA). Reaction intermediates were analyzed using liquid chromatography-mass spectrometry (LC-MS, Agilent 1290; Triple TOFTM 4600, AB Sciex, Framingham, MA, USA) in positive electrospray ionization (ES+) mode, with a capillary voltage of 4.5 kV, a cone voltage of 35 V, a source temperature of 50 °C, and a desolvation temperature of 25 °C. All electrochemical measurements were conducted at room temperature using a CHI 660E electrochemical workstation (Chenhua Co., Ltd., Shanghai, China).

3. Results and Discussion

3.1. Photocatalysts Comparison

The pure g-C3N4 synthesized in this study exhibited typical peaks at 2θ = 10° and 80° (Figure 2a). However, with the doping of Co, Ag, Mo, Bi and Zn, distinct peaks corresponding to CoO (JCPDS 42-1300), elemental Ag (JCPDS 04-0783), MoO3 (JCPDS 47-1320), elemental Bi (JCPDS, 44-1246), and ZnO (JCPDS, 36-1451) were observed in the respective catalysts, confirming the successful incorporation of metallic ions into the g-C3N4 structure. After 120 min of photocatalytic treatment, approximately 75.6% of 2,4-DCP was removed using Zn-rich g-C3N4, outperforming pure g-C3N4 and Co, Ag, Mo, and Bi-doped g-C3N4 catalyst (Figure 2b). The degradation kinetics followed a pseudo-first-order model, with regression coefficients greater than 0.99. The degradation rates for Co, Ag, Mo, and Bi-doped g-C3N4 ranged from 0.0046 to 0.0061 min−1, approximately half the rate achieved with Zn-doped g-C3N4 (Figure 2c), indicating a significant enhancement in photocatalytic activity due to Zn doping.
Furthermore, under light irradiation alone, only 20% of 2,4-DCP was removed (Figure 2d), which was about a quarter of the removal efficiency observed with Zn-rich g-C3N4 under light. However, it was five times higher than the removal by direct adsorption with Zn-rich g-C3N4 without light, indicating that photocatalytic degradation was the dominant mechanism for 2,4-DCP removal.

3.2. Photocatalytic Conditions Optimization

The photocatalytic conditions for Zn-rich g-C3N4 were optimized. As shown in Figure 3a, the degradation efficiency of 2,4-DCP increased by 5.7% across a salt concentration range of 0.1–2.3 wt.%, indicating that Zn-rich g-C3N4 maintained good stability in high-salinity wastewater. No significant changes in salt concentration were observed after treatment. Additionally, as the pH increased from 3 to 9 and 11, the degradation efficiency rose steadily from 70.1% to 84.6% and further to 99.4% (Figure 3b). After the reaction, the final pH stabilized between 6.0 and 6.6 (Figure 3c), although the equilibrium pH dropped during direct photodegradation. The underlying mechanism will be discussed in Section 3.4.
When the initial concentration of 2,4-DCP increased from 20 to 200 mg L−1, the degradation efficiency decreased from 100% to 51.9% (Figure 3d). Despite this, a significant removal of 103.8 mg L−1 2,4-DCP was achieved with an initial concentration of 200 m L−1 (Figure 3e). In the absence of the photocatalyst, direct adsorption of 2,4-DCP onto Zn-rich g-C3N4 resulted in concentrations of 2.1, 9, 12.2, and 13 mg L−1 with initial 2,4-DCP levels of 20, 50, 100, and 200 mg L−1, respectively (Figure 3f), corresponding to removal efficiencies of 2.1%, 1.8%, 1.2%, and 0.7%. These results demonstrated that photocatalytic degradation was the dominant mechanism for 2,4-DCP removal from saline wastewater.

3.3. Real Wastewater Treatment

The degradation of 2,4-DCP by Zn-rich g-C3N4 showed little variation between pure water and saline water (Figure 4a). In contrast, pure g-C3N4 achieved only 48.9% removal of 2,4-DCP in pure water and 38.6% in saline water. These results demonstrated the effectiveness of photocatalytic degradation of 2,4-DCP (or photocatalysis in the removal of 2,4-DCP).
Zn-rich g-C3N4 was also tested for the treatment of real saline organic wastewater (Figure 4b), which contained 2.2 wt.% salt and 385 mg L−1 total organic carbon (TOC). The major contaminants included 9.3 mg L−1 2,4-dichlorophenol, 13.2 mg L−1 2-chlorohydroquinone, 0.08 mg L−1 chloroacetophenone, and 0.17 mg L−1 2-chloropropionic acid. Under the optimized conditions, the removal efficiencies for these organics were 99.3%, 99.8%, 98.2%, and 99.9%, respectively, corresponding to a 67.7% reduction of TOC. In comparison, treatment with pure g-C3N4 achieved lower removal efficiencies of 41.2%, 44.1%, 32.5%, and 45.7%, with only 20.2% reduction of TOC. After treatment, the Zn-rich g-C3N4 was recovered and reused in subsequent cycles. Even after ten cycles, the degradation (or removal) efficiencies for 2,4-DCP and TOC remained at 95.5% and 56.5%, respectively, similar to the performance of the fresh catalyst (Figure 4c). This demonstrated the stability of Zn-rich Zn-rich g-C3N4 in the catalysis of recalcitrant organic compounds.

3.4. Zn-Rich g-C3N4 Characterization

The Zn-rich g-C3N4 primarily consisted of aggregated particles (Figure 5a) and retained sharp peaks characteristic of g-C3N4 at wavenumbers 1451, 1400, and 1316 cm−1 [18] (Figure 5b). However, the peaks at 1100 and 733 cm−1 shifted to lower values, and several peaks at 3080, 1240, and 805 cm−1 disappeared, indicating the incorporation of ZnO onto g-C3N4 via −NH and −OH bonds [21]. Additionally, the d(002) diffraction peak spacing of g-C3N4 changed from 0.324 to 0.317 nm, while the average crystalline size increased from 6.7 to 67.6 nm (Figure 2a). These changes suggested that Zn doping reduced the interlayer distance within the graphene-like structure and enhanced the stability of the crystal lattice [22].
The XPS spectra revealed two distinct peaks of Zn 2p at 1022.0 and 1045.1 eV [23] (Figure 5c). In the C 1s spectrum, the peaks at 288.4 and 284.8 eV became more intense (Figure 5d), indicating a greater exposure of −COOH and C−C sites due to Zn incorporation. Correspondingly, the N 1s peaks for the C−N=C bond at 398.7 eV increased (Figure 5e), while the peak for the N−(C)3 bond at 400.4 eV disappeared [18]. In the O 1s spectrum, a new peak at 533.4 eV was attributed to adsorbed O2 and/or water (Figure 5f). Collectively, these results demonstrated the successful doping of Zn into g-C3N4 and the corresponding changes in functional sites [21].
After the reaction, the collected Zn-rich g-C3N4 exhibited an irregular morphology with a smooth surface and sharp peaks corresponding to simoncolleite and ZnO (Figure 6a,b). Its XPS spectra also displayed typical peaks for Zn 2p and N 1s (Figure 6c,d). Notably, the relative percentage of the N peak and C=N−C peak in the N 1s spectrum was approximately 76%, which was similar to those of the fresh Zn-C3N4 particles.

3.5. Photocatalytic Mechanism Analysis

The removal of 2,4-DCP in the photocatalytic system involved three mechanisms: direct adsorption, direct photodegradation, and catalytic decomposition. Direct adsorption occurred on the surface sites of the catalyst; however, Zn-rich g-C3N4 had a lower total concentration of surface sites compared to pure g-C3N4. Since 2,4-DCP exists predominantly in its basic form (pKa = 7.8), it attaches to the catalyst surface through hydrogen bonds and van der Waals forces [24]. Consequently, only 3.97% of 2,4-DCP was removed via direct adsorption of Zn-rich g-C3N4, indicating a minimal contribution to the overall removal efficiency.
In addition to direct adsorption, photodegradation also played a significant role in 2,4-DCP disintegration. High-energy photons penetrated the water layer and directly attacked 2,4-DCP [25], leading to the formation of byproducts and contributing approximately 20% to the removal of 2,4-DCP from saline wastewater.
Upon initiating photocatalytic degradation, nearly 40% of 2,4-DCP was removed. The photocatalytic process involved four key steps: photon capture, charge excitation, electron-hole generation and transportation, and generation of free radicals. When pure g-C3N4 was introduced, it absorbed photons with wavelengths of less than 475 nm, causing electrons to transition from its valence band (1.34 eV) to the conduction band (−0.90 eV). This transition generated electron-hole pairs in the valence and conduction bands, respectively [26]. Once transported to the catalyst surface, the electrons were rapidly captured by ions, producing free radicals such as Cl·, ·OH, and·O2− [27]. This process initiated the photocatalytic degradation of 2,4-DCP.
Compared to pure g-C3N4, Zn-rich g-C3N4 exhibited superior photocatalytic performance. Doping Zn into g-C3N4 increased the band-gap energy from 2.74 eV to 3.35 eV while enhancing UV-vis absorbance in the range of 300–360 nm (Figure 7a,b). This indicated that Zn-rich g-C3N4 adsorbed more photons, which contributed to the generation of free radicals [28]. Furthermore, the Zn-rich g-C3N4 catalyst demonstrated a current intensity of 0.19 μA cm−2, 4.4 times greater than that of pure g-C3N4 (Figure 7c), significantly improved electron transfer efficiency due to Zn doping [29]. The Zn-rich g-C3N4 also exhibited lower charge transfer resistance, with a reference arc radius of 30.63 kΩ compared to g-C3N4, indicating its superior electron transfer capability (Figure 7d) [30].
The Tafel polarization curves (Figure 7e) revealed that the corrosion current of Zn-rich g-C3N4 was 9.61 × 10−6 A, an increase of one order of magnitude over pure g-C3N4, suggesting enhanced electron carrier properties. EPR spectra confirmed the generation of free radicals, including Cl· and ·OH (Figure 7f), upon the addition of Zn-rich g-C3N4 in saline water. The intensified diffraction peaks observed with Zn-rich g-C3N4 further suggested an increased yield of free radicals during photocatalytic degradation.
It was concluded that Zn doping enhanced the photocatalytic activity of g-C3N4 by improving UV-vis absorbance and electron transmission, resulting in greater formation of free radicals that positively influenced 2,4-DCP degradation.
Although Zn-rich g-C3N4 had fewer surface sites for the direct adsorption of 2,4-DCP, it exhibited excellent salt resistance in saline wastewater. In contrast, many other photocatalysts, including pure g-C3N4, had well-developed surface pores with sufficient sites for contaminant adsorption. Consequently, salt is often attached to these surface sites, significantly decreasing their capacity for the adsorption and oxidation of contaminants [21]. This phenomenon was detrimental to these photocatalysts, whereas Zn-rich g-C3N4 remained largely unaffected. In saline wastewater, chloride ions might react with photogenerated holes and free radicals, such as ·OH, to produce free chlorine radicals, further enhancing photocatalytic performance [31]. Therefore, the removal of 2,4-DCP primarily depended on the generation of free radicals rather than direct oxidation processes arising from direct adsorption.
Under the alkaline conditions, the availability of hydroxide ions increased. These ions attached to the surface of Zn-rich g-C3N4, leading to the production of hydroxyl radicals, which accelerated the degradation of 2,4-DCP and consumed free hydroxide ions simultaneously, resulting in a pH decrease. Additionally, the degradation products of 2,4-DCP consisted of various organic acids that released free H+ ions, further lowering the pH of the system.

3.6. Degradation Path of 2,4-DCP

2,4-DCP was disintegrated through two primary pathways (Figure 8). The first was direct photodegradation, where high-energy photons directly attacked the C−Cl bond of 2,4-DCP, leading to the formation of p-chlorophenol and o-chlorophenol (Figure S1a–c). Additionally, photons also targeted the C−O bond, resulting in phenolic aldehydes or the polymerization into dimers (Figure S1e,f). Several hours were usually needed to achieve significant results of 2,4-DCP degradation.
The second pathway involved photocatalytic degradation, where free radicals, such as ·OH and ·Cl, were generated on the catalyst’s surface and attacked adjacent 2,4-DCP molecules via the C−Cl bond, producing p-phenol and catechol. This process accelerated the dechlorination of 2,4-DCP, progressively increasing the concentration of free chloride ions in the treated water. As the free radicals continued to generate, they attacked the C=C and C−H bonds of primary byproducts, leading to the formation of secondary species, including maleic acid, fumaric acid, trihydroxyethylene, acetic acid, and oxalic acid (Figure S1g–n).

3.7. Prospectives

Zn-rich g-C3N4 demonstrated excellent photocatalytic performance in degrading organic contaminants in high-salt wastewater. In such environments, salt was toxic to catalysts, leading to its deterioration. The common practice is to wash the catalyst with deionized water to reactivate it. In this study, the degradation efficiency of 2,4-DCP slightly increased as salt concentrations increased from 0.1 to 2.3 wt.%. Notably, at a salt dosage of 6 wt.%, the degradation of 2,4-DCP reached 82.0%.
Although Zn-rich g-C3N4 exhibited low adsorption capacity for 2,4-DCP, indicating a limited number of surface adsorption sites, it demonstrated high salt tolerance. During the reaction, some dissolution of Zn into the wastewater occurred, but it remained minimal, even at an initial pH of 3. Importantly, the Zn-rich g-C3N4 maintained desirable reusability, with 2,4-DCP removal efficiency showing little variation after ten cycles. Thus, it sustained good photocatalytic performance even in high-salt wastewater.
Zn-rich g-C3N4 exhibited a broad absorbance spectrum, allowing it to effectively capture a wider range of photons across various energy levels. This enhancement in electron transmission contributed to a high photocatalytic degradation rate of 2,4-DCP, surpassing that of other synthesized catalysts such as g-C3N4-Cu2O, porous g-C3N4, O-g-C3N4, g-C3N4-Co-MOF, and CuPc/g-C3N4 (Table 1).
During the photocatalytic degradation process, the generated free radicals attacked 2,4-DCP and other organic contaminants, including rhodamine B, atrazine, and aniline, in a non-selective manner. Consequently, Zn-rich g-C3N4 holds promising potential for wastewater treatment across various industries, including paper, printing, dyeing, pharmaceuticals, and pickled food production [32,33].
Table 1. 2,4-DCP degradation by Zn-rich g-C3N4 in comparison with the other materials.
Table 1. 2,4-DCP degradation by Zn-rich g-C3N4 in comparison with the other materials.
No.LightCatalystExperimental ConditionRate
mg L−1 min−1		Ref.
1300 W HID0.2 g L−1 g-C3N4-Cu2OC4-CP = 1 mg L−1, 120 min0.008[34]
2300 W HID1.0 g L−1 porous g-C3N4C4-CP = 20 mg L−1, pH = 3, 120 min0.16[35]
31 mW cm−2 UV0.2 g L−1 O-g-C3N4C4-CP = 10 mg L−1, 50 μL H2O2, 30 min0.33[36]
450 mW cm−2 HID0.15 g L−1 g-C3N4-Co-MOFC4-CP = 20 mg L−1, 80 min0.12[37]
5300 W HID1.0 g L−1 CeO2/g-C3N4C4-CP = 5 mg L−1, 300 min0.008[38]
6HID2.5 g L−1 CuPc/g-C3N4C2,4-CP = 100 mg L−1, 240 min0.33[39]
7300 W HID1.0 g L−1 Pt/C3N4C5-CP = 20 mg L−1, 420 min0.05[40]
8300 W HID1.0 g L−1 g-C3N4/BiOIC4-CP = 10 mg L−1, 180 min0.01[41]
9300 W HID0.5 g L−1 NaxSCNNTsC4-CP = 20 mg L−1, 180 min0.10[42]
10HP 365 nm0.1 g L−1 Zn-rich g-C3N4C2,4-DCP = 100 mg L−1, 120 min0.63This work
HP: mercury lamp; HID: xenon lamp; UV: ultraviolet light.
Despite the advantages of Zn-rich g-C3N4 in the treatment of incalcitrant organic compounds in high-salt wastewater, further investigation is needed to synthesize Zn-rich g-C3N4 under milder conditions to reduce energy consumption. Additionally, the liquid reagents should be recycled, and the metallic chemicals should be extracted from spent waste [43].

4. Conclusions

This study investigated the synthesized Zn-rich g-C3N4 was used as a photocatalyst to degrade 2,4-DCP in a high-salt wastewater. While Zn-rich g-C3N4 demonstrated low adsorption capacity, indicating limited surface sites, it demonstrated high salt tolerance and broad light spectrum for absorbance, enabling efficient photon capture and generation of free radicals. EPR spectra confirmed the production of free radicals, including Cl· and OH. Additionally, the Zn-rich g-C3N4 photocatalysis system exhibited stability despite the variation of initial pH and 2,4-DCP concentrations. It effectively removed chlorophenols and total organic carbon (TOC) from real saline wastewater with a salinity of 2.2 wt.%. Product analysis led to the proposal of dehalogenation and ring-opening pathways of 2,4-DCP degradation. Overall, this study demonstrated the significant potential of Zn-rich g-C3N4 in the application of photocatalytic systems to treat incalcitrant organic compounds in high-salt wastewater.

Supplementary Materials

The following supporting information can be downloaded at: https://www.mdpi.com/article/10.3390/w16192756/s1, Text S1: Data calculation; Figure S1: Intermediates in photolysis system and Zn-type photocatalytic system.

Author Contributions

Conceptualization, S.Z.; Methodology, J.L.; Formal analysis, Y.Y.; Investigation, J.L.; Resources, X.W., L.W. and W.F.; Data curation, H.C.; Writing—original draft, H.C. and S.Z.; Project administration, Y.W., X.W. and L.W. All authors have read and agreed to the published version of the manuscript.

Funding

The authors acknowledge funding from the Young Scientist Fund (52100096), Shenzhen Science and Technology Innovation Commission (RCBS20210609104441072), the National Natural Science Foundation of China (52070038), the Science and Technology Program of Jilin Province (20240304153SF) and the Featured Innovation Project of Guangdong Provincial Department of Education (2023KTSCX050).

Data Availability Statement

Data is contained within the article.

Conflicts of Interest

The authors declare no conflict of interest.

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Water 16 02756 g001
Figure 1. Flow chart for preparing Me-C3N4 particles.
Figure 1. Flow chart for preparing Me-C3N4 particles.
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Figure 2. (a) XRD pattern of Me-C3N4, (b) photocatalytic degradation of 2,4-DCP by Me-C3N4, (c) the related kinetic constant, (d) photocatalytic degradation of 2,4-DCP with a different system. Conditions: 100 mg L−1 2,4-DCP, initial pH 5.4, 0.1 g L−1 catalyst dosage, 22 °C, 0.5 kW m−2 irradiance intensity.
Figure 2. (a) XRD pattern of Me-C3N4, (b) photocatalytic degradation of 2,4-DCP by Me-C3N4, (c) the related kinetic constant, (d) photocatalytic degradation of 2,4-DCP with a different system. Conditions: 100 mg L−1 2,4-DCP, initial pH 5.4, 0.1 g L−1 catalyst dosage, 22 °C, 0.5 kW m−2 irradiance intensity.
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Figure 3. Degradation of 2,4-DCP. (a) salt concentration, (b) initial pH, (c) final pH value, (d) initial concentration, (e) final concentration, and (f) direct adsorption. Conditions: 20–200 mg L−1 2,4-DCP, initial pH 3–11, 0.1 g L−1 Zn-C3N4 dosage, 22 °C, 0.5 kW m−2 irradiance intensity, 0.1–2.3 wt.% salinity.
Figure 3. Degradation of 2,4-DCP. (a) salt concentration, (b) initial pH, (c) final pH value, (d) initial concentration, (e) final concentration, and (f) direct adsorption. Conditions: 20–200 mg L−1 2,4-DCP, initial pH 3–11, 0.1 g L−1 Zn-C3N4 dosage, 22 °C, 0.5 kW m−2 irradiance intensity, 0.1–2.3 wt.% salinity.
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Figure 4. (a) Removal efficiency of 2,4-DCP in pure water and saline water, (b) removal efficiency of pesticide wastewater, and (c) ten continuous cycle experiments.
Figure 4. (a) Removal efficiency of 2,4-DCP in pure water and saline water, (b) removal efficiency of pesticide wastewater, and (c) ten continuous cycle experiments.
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Figure 5. (a) SEM image, (b) FTIR pattern, high resolution, (c) Zn 2p, (d) C 1s, (e) N 1s, and (f) O 1s XPS spectra.
Figure 5. (a) SEM image, (b) FTIR pattern, high resolution, (c) Zn 2p, (d) C 1s, (e) N 1s, and (f) O 1s XPS spectra.
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Figure 6. (a) SEM, (b) XRD, high resolution, (c) Zn 2p, and (d) N 1s XPS spectra of used Zn-C3N4.
Figure 6. (a) SEM, (b) XRD, high resolution, (c) Zn 2p, and (d) N 1s XPS spectra of used Zn-C3N4.
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Figure 7. (a) Kubelka-Munk spectra, (b) UV-vis spectra, (c) transient photocurrent response, (d) electrochemical impedance spectroscopy, (e) Tefel patterns, and (f) EPR spectra. Conditions: 100 mg L−1 2,4-DCP, initial pH 5.4, 0.1 g L−1 Zn-C3N4 dosage, 22 °C, 0.5 kW m−2 irradiance intensity, 1.2 wt.% salinity.
Figure 7. (a) Kubelka-Munk spectra, (b) UV-vis spectra, (c) transient photocurrent response, (d) electrochemical impedance spectroscopy, (e) Tefel patterns, and (f) EPR spectra. Conditions: 100 mg L−1 2,4-DCP, initial pH 5.4, 0.1 g L−1 Zn-C3N4 dosage, 22 °C, 0.5 kW m−2 irradiance intensity, 1.2 wt.% salinity.
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Figure 8. 2,4-DCP degradation pathway.
Figure 8. 2,4-DCP degradation pathway.
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Chen, H.; Wang, Y.; Zhu, S.; Wang, X.; Liu, J.; Wang, L.; Fan, W.; Yu, Y. High Salinity Tolerance of Zn-Rich g-C3N4 in the Photocatalytic Treatment of Chlorophenol Wastewater. Water 2024, 16, 2756. https://doi.org/10.3390/w16192756

AMA Style

Chen H, Wang Y, Zhu S, Wang X, Liu J, Wang L, Fan W, Yu Y. High Salinity Tolerance of Zn-Rich g-C3N4 in the Photocatalytic Treatment of Chlorophenol Wastewater. Water. 2024; 16(19):2756. https://doi.org/10.3390/w16192756

Chicago/Turabian Style

Chen, Hongyu, Ying Wang, Suiyi Zhu, Xiaoshu Wang, Jiancong Liu, Lei Wang, Wei Fan, and Yang Yu. 2024. "High Salinity Tolerance of Zn-Rich g-C3N4 in the Photocatalytic Treatment of Chlorophenol Wastewater" Water 16, no. 19: 2756. https://doi.org/10.3390/w16192756

APA Style

Chen, H., Wang, Y., Zhu, S., Wang, X., Liu, J., Wang, L., Fan, W., & Yu, Y. (2024). High Salinity Tolerance of Zn-Rich g-C3N4 in the Photocatalytic Treatment of Chlorophenol Wastewater. Water, 16(19), 2756. https://doi.org/10.3390/w16192756

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