Comparison of UV/PAA and VUV/PAA Processes for Eliminating Diethyl Phthalate in Water

Abstract

Diethyl phthalate (DEP) is a commonly utilized plasticizer that has gained significant attention due to its widespread occurrence in the environment and its harmful impact on human health. The primary objective of this study was to evaluate and compare several (ultraviolet) UV-(peracetic acid) PAA advanced oxidation processes based on hydroxyl radicals to degrade DEP. The effect of UV-LEDs incorporating PAA at different UV ranges (UV-A, λ = 365 nm; UV-C, λ = 254 nm and VUV, λ = 254 nm) was evaluated. The results demonstrated that DEP was successfully degraded in both the UVC/PAA (removal rate 98.28%) and VUV/PAA (removal rate 97.72%) processes compared to the UVA/PAA process (removal rate of 2.71%). The competitive method evaluated the contribution of R-O•, which were 24.08% and 33.92% in UVC/PAA and VUV/PAA processes, respectively. We also evaluated the effects of peroxymonosulfate (PMS) dosages, UV irradiation, pH and anion coexistence on the removal of DEP. In the UVC/PAA system, DEP degradation was particularly effective (removal rate about 95.52%) over a wider pH range (3–9). As the concentration of HCO₃⁻ ions increased, there may have been some inhibition of DEP removal. The inhibitory effect of HA and Cl⁻ ions on DEP removal were negligible. Analysis of the intermediates revealed that DEP degradation primarily occurred via two pathways: hydrolysis and hydroxylation reactions. This study presents a potential mnethod for the removal of phthalates and offers some guidance for the selection of appropriate disinfection technologies in drinking water treatment.

1. Introduction

Phthalates (PAEs) as a class of plasticizers are commonly employed in the manufacturing of toys, cosmetics, food packaging, and medical devices. With the rapid increase in use worldwide, PAEs can readily leach into the environment through point or non-point sources at various stages of manufacture, transportation and utilization. PAEs pollution has gradually become a hot topic of social concern, widely present in the atmosphere, water, soil, sediment, dust and other environmental media [1,2,3,4,5]. Ma et al. [6] detected total PAEs concentrations ranging from 930 to 2450 μg kg⁻¹ (dry weight (DW)) in soil and 790 to 3010 μg kg⁻¹ in vegetables in suburban Nanjing. Six PAEs were also detected in Chao Lake, China, in the range of 0.370–13.2 μg L⁻¹ by He et al. [7]. In the bohai rim region, concentrations of 16 PAEs ranging from 8.53 to 86.13 μg L⁻¹ were measured [8]. The presence of PAEs in the environment can cause a variety of hazards to humans and animals. PAEs can enter the human body through ingestion, respiratory inhalation and dermal contact, and their enrichment in the body may cause cytotoxicity, reproductive effects, endocrine effects and cancer [9,10,11]. Among all PAEs, diethyl phthalate (DEP) exhibits moderate to high ecological risk to aquatic plants, invertebrates, and vertebrates [8,12,13]. In particular, DEP is categorized as a high-priority contaminant by the U.S., the European Union and China [14]. Thus, it is important to remove DEP from aqueous environments, especially during the drinking water treatment.

A variety of methods have been used to treat DEP-containing surface water and wastewater, including adsorption, microbial degradation, base-catalyzed hydrolysis, advanced oxidation process (AOP) as well as catalytic ozonation [15,16,17,18,19,20,21]. Among various technologies, ultraviolet (UV) AOPs have attracted interest, because they have multiple functions for the control and disinfection of contaminants. UV/peracetic acid (PAA) has recently garnered growing interest as a disinfectant/oxidizer in water and wastewater treatment. PAA is considered an optimal disinfectant for drinking water treatment facilities due to its excellent disinfection effectiveness., low by-products, fast reaction, broad spectrum, environmental friendliness and simple process [22,23,24]. With the increasing requirements for water quality safety and environmental protection, the application prospect of PAA will be more broad [25,26,27].

In the UV/PAA system, the interaction of UV and PAA is synergistic because UV irradiation activates PAA, leading to the direct formation of highly reactive radicals, including •OH and carbon-centered radicals (R-C•; e.g., CH₃C(O)O• and CH₃C(O)OO•). Compared to UV/H₂O₂ system, PAA is more readliy activated because the peroxygen bond (O-O) is slightly longer (1.443 Å), and the O-O bond energy is lower (170 kJ mol⁻¹) in PAA compared to H₂O₂ (1.427 Å and 210 kJ mol⁻¹) [28]. Although the R-O• is selective for contaminants, it has a longer lifetime and higher concentration in the water, which could contribute more to the degradation of pollutants. Among UV/PAA processes, there are three wavelengths of UV lamps (e.g., UVA, UVC and VUV) [26,29,30]. UVA (wavelengths of 320–380 nm) has a lower energy compared to conventional UVC (wavelength 254 nm), resulting in less absorbance in the water matrix [31]. The interaction of VUV (wavelength 185 nm) and dissolved oxygen produces hydrogen atoms (H•), solvated electrons (e_aq⁻) and •OH [32]. The selection of the appropriate UV wavelength is directly related to the activation of PAA and the degradation efficiency of DEP, which is indirectly related to the cost and effectiveness of the water treatment process. However, information on the degradation of DEP through the UV/PAA process is very limited and the associated mechanism remain unknown.

The objectives of this study were to systematically examine the kinetics of DEP degradation in UVA/PAA, UVC/PAA, and VUV/PAA systems. The impact of operating parameters such as PAA dosage, initial pH, and common anions on DEP removal was also assessed. Additionally, the potential catalytic mechanism and the roles of •OH and ROS in contaminant degradation were investigated through quenching experiments, electron paramagnetic resonance (EPR) analysis, and the probe method. Furthermore, an electrical energy per order (EE/O) analysis and a cost evaluation of the UVC/PAA process were conducted. Lastly, the potential degradation pathways of DEP were proposed based on gas chromatography-mass spectrometry (GC-MS) results.

2. Materials and Methods

2.1. Materials

Diethyl phthalate (DEP, purities > 99%), sodium thiosulfate (Na₂SO₃), and acetic acid (CH₃COOH,97% w/w) were purchased from Aladdin (Shanghai, China). H₂O₂ (30% w/w) solution, para-chlorobenzoic acid (pCBA, 99% purity), sulfuric acid (H₂SO₄, 96%) and sodium hydroxide (NaOH) were purchased from Sinopharm Chemical Regent Co., Ltd. (Shanghai, China). Sodium chloride (NaCl), sodium hydrogen carbonate (NaHCO₃), humic acid (HA, >99% purity), sodium nitrate (NaNO₃) and potassium permanganate (KMnO₄) were purchased from Maclean Biochemical Technology Corporation (Shanghai, China). All the solutions were prepared using Milli-Q water.

The PAA stock solution was prepared by mixing acetic acid (97% w/w) and H₂O₂ (30% w/w) in a 1:1 ratio by volume, followed by the addition of sulfuric acid (97% w/w) to 3% w/w. The concentration of PAA in the stock solution ranged from 18% to 20% by weight, and the concentration of H₂O₂ ranged from 8% to 10% by weight. The PAA concentration was calibrated once a week. The PAA and H₂O₂ concentrations were determined using a two-step titration. The total peroxide in the solution was determined by indirect iodometric method, and the H₂O₂ in the solution was calibrated with 0.1 M KMnO₄ standard solution, the difference between the total peroxide and the H₂O₂ was the PAA concentration in the solution.

2.2. Experimental Procedures

All degradation experiments were performed in a photochemically operated reactor equipped with different UV lamps and magnetic stirrer. PAA was added to a 100 mL reaction solution containing DEP (2.0 μM) and reacted under UV irradiation (10 W low pressure mercury light: UVA (λ = 365 nm), UVC (λ = 254 nm) and VUV (λ = 185 nm)). 1 mL aliquots were sampled at predetermined time (5 min) intervals. Subsequently, excess Na₂S₂O₃ was added to quench the residual oxidant for further analysis. The designed initial pH of the reaction solution was adjusted with 1 M H₂SO₄ or 1 M NaOH. The initial pH and UV intensity of the DEP solution at the time of the experiment were 9.0 and 855 μW/cm², respectively. Degradation experiments were carried out under these conditions for 30 min. To investigate the effect of aqueous matrix on the degradation of DEP by the UV/PAA process, HCO₃⁻ (0–5.0 mM), NO₃⁻ (0–5.0 mM), Cl⁻ (0–5.0 mM) and HA (0–5.0 mg/L) were added to the above reaction solutions, respectively. Tert-Butanol (TBA) and methanol (MeOH) were added to the reaction solution at the start to investigate the involvement of free radicals in DEP degradation via the UV/PAA process. All experiments were conducted in triplicate.

The concentrations of DEP and pCBA were determined using an Agilent 1100 series HPLC equipped with a Poroshell 120 EC-C18 column (4.6 mm × 50 mm, 2.7 μm, Agilent, CA, USA). The column temperature was maintained at 40 °C when measuring both substances. The mobile phase used to analyze the DEP concentration consisted of 70% methanol, 30% water. The column flow rate was set to 1.0 mL/min and the detection wavelength was 228 nm. The mobile phase was used to analyze the concentration of pCBA by 90% methanol and 10% acetic acid. The column flow rate was set to 1 mL/min and the detection wavelength was 23 nm. The intermediates were determined by gas chromatograohy-mass spectrometry (GC-MS, Agilent 6890N gas chromatograph) equipped with FID detector (HP-5 column). The temperature of inlet and detector were 250 °C and 300 °C, respectively. First, the temperature was increased to 150 °C for 3 min at a rate of 15 °C/min, and then to 300 °C for 3 min.

In this study, pCBA was used as a •OH probe compound to assess the contribution of reactive radicals. The contribution of these radicals to DEP degradation in the UV/PAA process can be calculated using Equations (1)–(4). k₁ and k₂ represent the second-order rate constants for the reaction of •OH with pCBA and DEP, respectively (k₁ = 5 × 10⁹ M⁻¹ s⁻¹ [33], k₂ = (3.7 ± 0.1) × 10⁹ M⁻¹ s⁻¹ [34]. k₃ is the second-order rate constant for the reaction of RC• with DEP. k_obs•OH and k_{obs RC•} are the rate constants for DEP degradation by •OH oxidation and RC• oxidation, respectively.

$$k_{\mathrm{obs}}=\ln {\displaystyle \frac{\left|\mathrm{DEP}\right|}{{\left|\mathrm{DEP}\right|}_0}}/\mathrm{t}$$

(1)

$$k_{\mathrm{obs}}{=k}_{\mathrm{obs}•\mathrm{OH}}{+k}_{\mathrm{obs},\mathrm{R}-\mathrm{O}•}$$

(2)

$$\ln {\displaystyle \frac{\left|\mathrm{pCBA}\right|}{{\left|\mathrm{pCBA}\right|}_0}}{/\mathrm{t}=k}_1\int {[•\mathrm{OH}]}_{\mathrm{ss}}\mathrm{dt}$$

(3)

$$\ln {\displaystyle \frac{\left|\mathrm{DEP}\right|}{{\left|\mathrm{DEP}\right|}_0}}{/\mathrm{t}=k}_2\int {[•\mathrm{OH}]}_{\mathrm{ss}}{\mathrm{dt}+k}_3\int {[\mathrm{R}-\mathrm{O}•]}_{\mathrm{ss}}\mathrm{dt}$$

(4)

The electrical energy per order, EE/O has been widely used to assess the performance of UV systems utilizing low-pressure mercury lamps. EE/O is defined as the amount of electric energy (in kWh) required to reduce the concentration of a contaminant by one order of magnitude (90% removal) in 1 m³ of water. It is calculated according to Equations (5)–(7) [35,36]. where P is the rated power or energy input of the lamp system (0.01 kW), V (L) is the solution volume, k is the rate constant of pollutant degradation (min⁻¹), C₀ and C_t are the concentrations of DEP at reaction times 0 and t, r₁ is the consumption of PAA (mg L⁻¹) and E_PAA is the equivalent electric energy consumption to produce a milligram of PAA (40 kWh kg⁻¹).

$$\mathrm{EE}/\mathrm{O}={\displaystyle \frac{1000\mathrm{Pt}}{60\mathrm{V} \log (\frac{{\mathrm{C}}_0}{{\mathrm{C}}_{\mathrm{t}}})}}={\displaystyle \frac{38.38\mathrm{P}}{\mathrm{V}k}}$$

(5)

$${\mathrm{EE}/\mathrm{O}}_{\mathrm{UV} }={\displaystyle \frac{1000\mathrm{Pt}}{60\mathrm{V} \log (\frac{{\mathrm{C}}_0}{{\mathrm{C}}_{\mathrm{t}}})}}={\displaystyle \frac{38.38\mathrm{P}}{{\mathrm{V}k}_{\mathrm{UV}}}}$$

(6)

$${\mathrm{EE}/\mathrm{O}}_{\mathrm{UV}/\mathrm{PAA}}={\displaystyle \frac{38.38\mathrm{P}}{{\mathrm{V}k}_{\mathrm{UV}/\mathrm{PAA}}}}+{\displaystyle \frac{{\mathrm{r}}_1{ \mathrm{E}}_{\mathrm{PAA}}}{1000}}$$

(7)

3. Results and Discussion

3.1. Effect of UVA, UVC and VUV on the Removal of DEP

The performance of DEP degradation was compared on the different oxidation processes including UVA irradiation, UVC irradiation, VUV irradiation, PAA alone, UVA/PAA, UVC/PAA, UVC/PAA and VUV/PAA. As depicted in Figure 1a, UV irradiation and PAA alone were ineffective in degrading DEP. This phenomenon may be attributed to the low molar absorption coefficients and quantum yields under UVA, UVC and VUV irradiation. Besides, the removal of DEP by UVA/PAA system is negligible. Since UV wavelength required for PAA activation is less than 340 nm, UVA is not able to activate PAA, which was consistent with the result of Orlando and Tyndall [37]. The degradation of DEP followed pseudo-first-oder kinetics. Compared to UVA/PAA system, both the UVC/PAA system (98.28% DEP removal in 30 min, with an apparent first-order rate constant (k_obs) of 0.134 min⁻¹) and the VUV/PAA system (97.12% DEP removal in 30 min, with an apparent first-order rate constant (k_obs) of 0.120 min⁻¹) significantly enhanced DEP degradation (Figure 1b). Na et al. [38] studied the removal of DEP from aqueous solutions using a combination of ultrasonic (US) and ultraviolet (UVC) irradiation with titanium dioxide (TiO₂), achieving approximately 60% removal in 120 min at 283 kHz. Wang et al. [34] investigated that the UV/peroxynitrite system could achieve a high DEP degradation efficiency (92.6%) within 60 min at a peroxynitrite dose of 200 μM. In conclusion, the UV/PAA process can better degrade DEP effectively in a short time. The high degradation rates of DEP in UVC/PAA and VUV/PAA systems were attributed to the generation of •OH and ROS by the 185 nm photons and 254 nm photons (Equations (8)–(14)).

$${\mathrm{CH}}_3{\mathrm{CO}}_3\mathrm{H}+\mathrm{photosis}\to ·{\mathrm{CH}}_3{\mathrm{CO}}_2+·\mathrm{OH}$$

(8)

$$·{\mathrm{CH}}_3{\mathrm{CO}}_2\to ·{\mathrm{CH}}_3{\mathrm{O}}_2$$

(9)

$$·{\mathrm{CH}}_3{+\mathrm{O}}_2\to ·{\mathrm{CH}}_3{\mathrm{O}}_2$$

(10)

$${\mathrm{CH}}_3{\mathrm{CO}}_3\mathrm{H}+·\mathrm{OH}\to ·{\mathrm{CH}}_3{\mathrm{CO}}_3{+\mathrm{H}}_2\mathrm{O}$$

(11)

$${\mathrm{CH}}_3{\mathrm{CO}}_3\mathrm{H}+·\mathrm{OH}\to ·{\mathrm{CH}}_3{\mathrm{CO}+\mathrm{O}}_2{+\mathrm{H}}_2\mathrm{O}$$

(12)

$${\mathrm{CH}}_3{\mathrm{CO}}_3\mathrm{H}+·\mathrm{OH}\to ·{\mathrm{CH}}_3\mathrm{COOH}+·\mathrm{OOH}$$

(13)

$${\mathrm{CH}}_3{\mathrm{CO}}_3\mathrm{H}+·{\mathrm{CH}}_3{\mathrm{CO}}_2\to ·{\mathrm{CH}}_3{\mathrm{CO}}_3{+\mathrm{CH}}_3\mathrm{COOH}$$

(14)

Due to the presence of H₂O₂ in PAA solution, the contribution of H₂O₂ in UVC/PAA and VUV/PAA systems were investigated (the utilized H₂O₂ dosage was equal to the H₂O₂ concentration in 0.5 mM PAA solution). As can be seen from Figure 2a,b and Tables S1 and S2, the removal of DEP was 87.86% and 83.56% in the UVC/H₂O₂ (k_obs = 0.069 min⁻¹) and VUV/H₂O₂ (k_obs = 0.057 min⁻¹) systems, respectively. The reaction rate constants of H₂O₂ were lower than those of PAA because the molar absorption coefficient (absorbance value at a certain wavelength with a concentration of 1 M and an optical range of 1 cm) of PAA was 18.45 M⁻¹ cm⁻¹ at 254 nm with a quantum yield of 1.20 mol/Einstein, while the molar absorption coefficient of H₂O₂ was 18.7 M⁻¹ cm⁻¹ at 254 nm with a quantum yield of 1.0 mol/Einstein. This result also reveals that the role of PAA in enhancing DEP removal in the UV/PAA system.

The potential reactive oxygen species (ROS) involved in DEP degradation were further identified through EPR analysis, using DMPO as a spin trap for •OH and R-O• radicals. As shown in Figure S2, the specific EPR signals of DMPO •OH (1:2:2:1) and R-O• were observed in the UV/PAA system. Thus, the degradation of DEP is related to the free radicals generated in the system. The reactive radicals scavenging assay was applied to explore the importance of reactive radicals to DEP removal. TBA was effective in quenching •OH (second-order rate constant was (3.8–7.6) × 10⁸ M⁻¹ s⁻¹) but had no effect on R-O•. The degradation rate of DEP decreased from 0.134 min⁻¹ to 0.0145 min⁻¹ and from 0.120 min⁻¹ to 0.028 min⁻¹ after the addition of excess TBA (10 mM) to the UVC/PAA and VUV/PAA systems, respectively. Therefore, DEP degradation in UVC and UVU systems is mainly dependent on •OH. MeOH containing α-H effectively quenched R-O• and •OH. The addition of MeOH effectively inhibited DEP degradation both in UVC/PAA and VUV/PAA systems, and the k_obs of DEP were reduced to 0.0091 min⁻¹ and 0.0090 min⁻¹, respectively.

The probe method was utilized to assess the contribution of reactive radicals in these systems. pCBA was usually acted as a •OH probe compound [39,40,41]. When TBA was present, 5 μM of pCBA showed little degradation in the UVC/PAA and VUV/PAA systems over 30 min, suggesting that pCBA cannot be scavenged by R-O•. Figure S3 show the competitive degradation of pCBA and DEP during UVC/PAA and VUV/PAA systems. The k_obs of p-CBA were 0.087 min⁻¹ and 0.076 min⁻¹ in UVC/PAA and VUV/PAA systems, respectively. The steady-state concentration of •OH ([•OH]_SS) and R-O• ([R-O•]_SS) were calculated by competitive degradation of pCBA during UVC/PAA and VUV/PAA process (Equations (9)–(11)). The contribution of •OH to DEP degradation were 75.92% and 66.08% in the UVC/PAA and VUV/PAA systems, respectively (Figure 2c,d). It was also demonstrated that the •OH in the UVC/PAA and VUV/PAA systems played a crucial role in DEP degradation. However, the slightly lower •OH radical contribution in the VUV/PAA system was due to the production of hydrogen atoms (H•) and solvated electrons (e_aq⁻) under VUV irradiation, which was capable to remove a certain amount of DEP.

3.2. Factors Affecting DEP Degradation

From the above experiments, it can be seen that the UVC/PAA system degraded DEP slightly better than the VUV/PAA system. Moreover, VUV generates a certain amount of ozone under light, which is harmful to the environment. Furthermore, 254 nm UV lamps is most widely used in the practical application. The following experiments were all conducted in the UVC/PAA system.

The efficiency of DEP degradation was further investigated at different PAA dosages (ranging from 0.125 to 1.0 mM). As shown in Figure 3a,b, the performance of DEP degradation was enhanced from 75.6% to 99.9% with the increasing of PAA concentrations from 0.125 mM to 1.0 mM, and the k_obs of DEP increased from 0.048 min⁻¹ (linear correlation coefficient (R²) = 0.980) to 0.230 min⁻¹ (R² = 0.948) (Table S3). Moreover, PAA dosage was linearly proportional to the value of k_obs. This is due to the fact that more reactive substances (e.g., •OH and R-O•) were produced at higher PAA concentrations. When the PAA concentration was 0.5 mM, the DEP was all degraded within 30 min, which was more economical.

An increase in UV intensity can increase the number of photons that can be provided by the light source per unit time [42]. The detailed data of k_obs were given in Table S4. In the UV/PAA system, the enhancement of UV intensity leads to the increase of reactive radicals generation (e.g., •OH and R-O•), which are essential for the degradation of DEP. The removal of DEP was conducted at UVC light intensities of 378, 615 and 855 μW/cm² (Figure 3c,d). The result shows that DEP removal after 30 min of reaction was 42.83%, 80.66% and 98.28%, respectively, which are consistent with primary reaction kinetics.

Since the change of solution pH can alter the speciation of PAA oxidant, the effect of different initial pH (3.0–11.0) on the degradation of DEP was assessed in the study. As shown in Figure 3e,f and Table S5, the corresponding k_obs values were 0.182 min⁻¹, 0.210 min⁻¹, 0.100 min⁻¹, 0.134 min⁻¹ and 0.048 min⁻¹ as pH increased from 3.0 to 11.0. Overall, the acidic conditions were favorable for the PAA oxidation system. The effect of initial pH on DEP can be explained by differences in PAA status at different pH values. The two forms of PAA, conjugate acid (PAA⁰) and conjugate base (PAA⁻), are presented at different pH. Because pK_a value of PAA is 8.2, PAA⁰ is the predominant form in acidic and slightly alkaline solution (pH = 3.0, 5.0 and 7.0), and PAA⁻ is the predominant form in alkaline solution (pH = 9.0 and 11.0). In alkaline solution, the faster photolysis and higher quantum yield of PAA⁻ could lead to an increase in -OH and R-O- concentrations comparing to PAA. However, PAA⁻ is more capable of scavenging •OH than PAA (k_•OH, PAA⁻ and k_•OH, PAA⁰ were (9.97 ± 2.30) × 10⁹ M⁻¹ s⁻¹ and (9.33 ± 0.30) × 10⁸ M⁻¹ s⁻¹, respectively) [43,44]. Moreover, the •OH redox potential was lower in alkaline pH conditions than in acidic or neutral pH conditions, explaining that the reaction rate was lower in alkaline solution than in acidic solution. This result also illustrated the dominance of •OH rather than R-O• in the UVC/PAA system.

3.3. Possible Degradation Pathways

To gain a better understanding of the DEP and PAA degradation process under UVC irradiation, the degradation products of DEP were analyzed using GC-MS, with the detailed results presented in Figure S4. Ten intermediates were identified in the UVC/PAA system as shown in Figure 4 (P1–P14), and four potential degradation pathways were proposed based on the identified degradation products. In pathway I and II, DEP undergoes a hydrolysis reaction to cleave the ester group to form the monoester compound (P1 and P2) [45]. Then, P1 and P2 are further hydrolyzed to form P5 and P6. P7 is obtained by α-cleavage of the C-C bond connecting the C-C=O molecule to the aromatic ring [46]. P7 is further hydroxylated to form P8 with the presence of free radicals. In Pathway III and Pathway IV, P3 and P4 are formed by hydroxylation and acylation during DEP degradation. Then P3 can be further hydrolyzed to form P5, P6 and P7. P4 can be hydrolyzed to form P3 and P9, and then P9 can continue to be hydrolyzed to form P10. Finally, these aliphatic acids are mineralized into CO₂ and H₂O.

3.4. Effect of Anions on DEP Degradation

Anions and natural organic matter (NOM) have a strong influence on contaminant degradation in AOP processes. Cl⁻, HCO₃⁻ and NO₃⁻ are three common anions, while HA is the prevalent NOM in real water. Therefore, the effects of the four aqueous matrices on the degradation of DEP in the UV/PAA system were further investigated.

As shown in Figure 5a and Table S6, the degradation rate of DEP was slightly increased from 0.134 min⁻¹ to 0.168 min⁻¹ with increasing Cl⁻ concentration from 0 to 1.0 mM. Hu et al. [47] have observed similar trends. However, as the Cl⁻ concentration increased from 0.1 to 5.0 mM, the k_obs of DEP degradation gradually decreased from 0.134 min⁻¹ to 0.128 min⁻¹. This phenomenon can be explained by the fact that the presence of Cl⁻ at lower concentrations could scavenges R-C• to produce Cl• as described in Equation (15). The Cl• further reacted with H₂O₂ to form HOCl•⁻ (Equation (16)). The rate constant of the HOCl•⁻ dissociation reaction is higher than the opposite rate constant, thus HOCl•⁻ tends to dissociate to form •OH in Equation (17), which enhances the degradation of DEP. As the concentration of Cl⁻ continues to increase, Cl⁻ can scavenge •OH to form HOCl•⁻. When the Cl⁻ concentration was increased to 5.0 mM, Cl⁻ scavenges •OH and further converted to HOCl⁻• and other active chlorine species in Equation (17) [39,48]. The decrease in •OH and the low reactivity of the chlorine-containing radicals to DEP may be responsible for the decrease in degradation rate.

$$·{\mathrm{CH}}_3{\mathrm{CO}}_3{+\mathrm{Cl}}^{-}{+\mathrm{H}}^{+}\to \mathrm{Cl}·{+\mathrm{CH}}_3{\mathrm{CO}}_3\mathrm{H}$$

(15)

$$\mathrm{Cl}·{+\mathrm{H}}_2\mathrm{O}\leftrightarrow \mathrm{HOCl}{·}^{-}{+\mathrm{H}}^{+}$$

(16)

$$\mathrm{HOCl}{·}^{-}\leftrightarrow ·{\mathrm{OH}+\mathrm{Cl}}^{-}$$

(17)

The effect of HCO₃⁻ on DEP degradation in the UV/PAA system was shown in Figure 5b. The concentration range of HCO₃⁻ (0–5.0 mM) did not expressively promote DEP degradation in the UV/PAA system. It can be found that the k_obs of DEP degradation decreased from 0.134 min⁻¹ to 0.047 min⁻¹ as the HCO₃⁻ concentration increased from 0 mM to 5.0 mM (detailed data of k_obs are given in Table S7). This is also similar to the conclusion of wang et al. [34]. It is well known that HCO₃⁻ and CO₃²⁻ can scavenge •OH to produce CO₃•⁻ at the rate of 8.5 × 10⁶ M⁻¹ s⁻¹ and 3.9 × 10⁶ M⁻¹ s⁻¹ in Equations (18) and (19). Then CO₃•⁻ could further react with PAA to produce R-C• at the rate of 3.9 × 10⁸ M⁻¹ s⁻¹ (Equation (20)) [39]. Since the reactivity of CO₃•⁻ and R-C• for DEP are lower than that of •OH, the degradation of DEP is inhibited with the presence of high HCO₃⁻ concentration.

$$·{\mathrm{OH}+\mathrm{HCO}}_3^{-}\to {\mathrm{CO}}_3{·}^{-}{+\mathrm{H}}_2\mathrm{O}$$

(18)

$$·{\mathrm{OH}+\mathrm{CO}}_3^{2-}\to {\mathrm{CO}}_3{·}^{-}{+\mathrm{OH}}^{-}$$

(19)

$${\mathrm{CO}}_3{·}^{-}{+\mathrm{CH}}_3{\mathrm{CO}}_3\mathrm{H}\to ·{\mathrm{CH}}_3{\mathrm{CO}}_3{+\mathrm{HCO}}_3^{-}$$

(20)

The effect of different concentrations of NO₃⁻ on the degradation of DEP in the UV/PAA system was also investigated and the results are shown in Figure 5c and Table S8. The degradation of DEP increased slightly (k_obs from 0.134 min⁻¹ to 0.174 min⁻¹) as the NO₃⁻ concentration increased from 0 to 0.1 mM. NO₃⁻ is a common photosensitizer that can be excited under UV irradiation to generate NO•, NO₂• and O•⁻ in Equations (21)–(23), which further form •OH in Equation (24) [47]. The promotion of DEP degradation in low concentrations of NO₃⁻ may be related to the fact that the positive effect can compensate the negative effect. However, NO₃⁻ had an inhibition effect (k_obs decreased from 0.174 min⁻¹ to 0.157 min⁻¹) when the concentration of NO₃⁻ was increased to 1.0 mM. This may be due to the fact that NO₃⁻ can compete with PAA for UV absorption and NO₃⁻ can also be converted to NO₂⁻ upon UV irradiation in Equation (25) [49].

$${\mathrm{NO}}_3^{-}+\mathrm{hv}\to {\mathrm{NO}}_2·+\mathrm{O}{·}^{-}$$

(21)

$${\mathrm{NO}}_2^{-}+\mathrm{hv}\to \mathrm{NO}·+\mathrm{O}{·}^{-}$$

(22)

$${\mathrm{NO}}_3^{-}+\mathrm{hv}\to {\mathrm{NO}}_2^{-}+0{.5\mathrm{O}}_2$$

(23)

$$\mathrm{O}{·}^{-}{+\mathrm{H}}_2\mathrm{O}\to ·{\mathrm{OH}+\mathrm{OH}}^{-}$$

(24)

$${\mathrm{NO}}_2^{-}+·\mathrm{OH}\to {\mathrm{NO}}_2·{+\mathrm{OH}}^{-}$$

(25)

The effect of HA concentration on DEP degradation in the range of 0–5.0 mg/L was shown in Figure 5d and Table S9. The degradation of DEP was slightly inhibited. HA, as a photosensitizer, can compete with DEP and PAA for UV photon adsorption, thus inhibiting the formation of reactive radicals during UV/PAA process. HA scavenges •OH with a second-order rate constant of 2.5 × 10⁴ L mg⁻¹ s⁻¹ [48]. A recent study by Chen et al. [39] has shown that HA can also scavenge R-C in the UV/PAA system. However, when the HA concentration increased from 1.0 mg/L to 5.0 mg/L, HA slightly promoted DEP degradation. This may be due to the fact that HA can be excited by UV irradiation to produce photogenerated reactive substances, including •OH, single-linear oxygen (¹O₂), and so on. These substances reacted with DEP to promote its degradation. As the concentration of HA increased to 5 mg/L, high concentrations of HA can act as a free radical scavenger, thus reducing the removal efficiency of DEP.

3.5. Energy Consumption

Since UV/PAA degradation is an electricity-intensive process, and electricity can represent a significant portion of operating costs. Therefore, the electrical energy per order (EE/O) was evaluated for the UVA/PAA, UVC/PAA and VUV/PAA processes to identify the energy consumption during the DEP degradation. The EE/O values were calculated in the UVC/PAA and VUV/PAA systems in Figure 6. In the UV based system, the energy consumption was mainly from UV irradiation. The VUV/PAA system consumed more energy than the UVC/PAA system with EE/O values of 33.24 and 29.84 kWh m⁻³ respectively. The UVC/PAA system reduced the energy consumption by 10.23% as compared to the VUV/PAA system.

4. Conclusions

In this study, the degradation of DEP was comprehensive evaluated in UVA/PAA, UVC/PAA and VUV/PAA systems. The UVC/PAA and VUV/PAA processes showed significant synergistic effects in degrading DEP with degradation efficiencies of 98.28% and 97.12% within 30 min. Effective degradation of DEP was also observed in UVC/PAA system over a wide pH range of 3.0–9.0. The free radicals in the UV/PAA system were identified by EPR, quenching assay and probe assay, where •OH was the main reactive free radical for the degradation of DEP. The competitive method evaluated the contribution of R-O•, which were 24.08% and 33.92% in UVC/PAA and VUV/PAA respectively. The presence of 5.0 mM HCO₃⁻ resulted in a degradation Of DEP from 98.28% to77.69%. Moreover, the EE/O also indicated that the UVC/PAA system was an energy-efficient and cost-effective method for water treatment. Thes results provided valuable information and reference for the effective removal of DEP from drinking water treatment plants. The potential toxicity and persistence of by-products should be further investigated in the future, as well as exploring energy-saving strategies to utilize solar-assisted UV systems to bring the process closer to practical applications.