Photocatalytic Degradation of Trichloroethylene Under Different Environmental Conditions: Kinetics and Carbon Isotope Effects

Abstract

Photocatalytic degradation technology is an important tool for treating trichloroethylene (TCE) pollution in water bodies. While previous studies have focused on catalyst optimization and degradation kinetics for trichloroethylene (TCE) photocatalysis, the systematic impact of environmental conditions on carbon isotope effects (ε) and their mechanistic implications remains poorly understood. This limits the reliability of quantitative isotope analyses in field applications. We conducted a series of laboratory experiments on the photocatalytic degradation of TCE to investigate the kinetic and isotopic effects under different conditions. Increasing the concentration of TCE, anions (NO₃⁻ and HCO₃⁻), and cations (Mg²⁺ and Ca²⁺) decreased the photocatalytic degradation of TCE. O₂ will increase the degradation efficiency of TCE. The dose required to achieve maximum photocatalytic efficiency varies for different types of catalysts, which needs to be determined on a case-by-case basis. TCE photocatalytic degradation had a small carbon isotope effect (ε = −2.0 ± 0.2‰ to −3.2 ± 0.5‰), which was slightly affected by the catalyst dosage and species (TiO₂ and ZnO), and concentrations of TCE, O₂, and inorganic ions (NO₃⁻, HCO₃⁻, Mg²⁺, and Ca²⁺). The ε values are stable and reproducible and relatively insensitive to our selected environmental factors in this study, which can reduce the uncertainty of applying stable carbon isotope enrichment factors to quantify the photocatalytic reaction for remediation of TCE contaminated sites.

1. Introduction

Trichloroethylene (TCE) contamination arises primarily from emissions, spills, or volatilization of solvents into environmental media during industrial cleaning, metal degreasing, dry cleaning, and electronics manufacturing [1]. Due to its resistance to biotic and abiotic degradation under natural conditions, TCE is one of the most widespread organic contaminants in soil and water [2,3,4]. TCE and its degradation intermediates pose a significant threat to human health as carcinogens [5,6,7]. Several methods have been reported for the degradation of TCE in soil and water, including physical [8,9,10], chemical [11,12,13,14,15], and biological techniques [16,17,18]. Physical adsorption achieves rapid removal of pollutants through the retention of pollutants on the surface of the material, which is easy to operate but unable to degrade the pollutants, and the adsorbent is easily saturated. Chemical oxidation (e.g., the Fenton method) utilizes a strong oxidant to efficiently mineralize the pollutants, but at a high cost, high risk of by-products, and harsh reaction conditions. Biological degradation relies on the decomposition of pollutants by microorganisms, which is environmentally friendly but takes a long period of time and is subject to environmental constraints. In contrast, a photocatalytic reaction is an environmentally friendly advanced oxidation process (AOP) that can completely degrade organic pollutants and avoid secondary pollution, at a lower cost than oxidation with chemical reagents and with an efficiency far exceeding biodegradation [19].

ZnO and TiO₂ are two of the most effective catalysts for the removal of organic chemicals [20,21]. Their catalytic mechanism is very similar. Under UV light irradiation, electrons and holes are generated in the conduction band (CB) and valence band (VB), respectively, forming electron–hole pairs. The excited electrons can react with oxygen molecules on the surface of the catalyst or in water to produce O₂•⁻, and the generated holes oxidize the H₂O and OH⁻ to produce •OH [22]. The •OH and O₂•⁻ have strong oxidizing abilities and can decompose TCE into harmless or less harmful compounds [23], with chloride ions and CO₂ as the main products, and several reaction intermediates (e.g., formic acid, dichloroacetic acid, and monochloroacetic acid) [24]. TiO₂ (3.2 eV) and ZnO (3.37 eV) have similar band gap energies, and their photocatalytic abilities are similar. In addition, ZnO catalysts are relatively cheaper than TiO₂ catalysts, making the use of ZnO more economical for large-scale water purification processes [21]. However, the main disadvantage of ZnO is its wide band gap energy; therefore, the absorption of light by ZnO is limited to the UV range. Moreover, rapid recombination of light-generated charges may occur, which reduces the photocatalytic performance of ZnO [25,26].

Stable carbon isotope analysis is a powerful tool for the assessment of remediation performance at sites with organic contaminants [27]. It is primarily based on carbon isotope effects during processes such as pollutant degradation and transformation (Equation (1)). This method has been successfully used to assess the remediation of contaminated water bodies [28,29,30].

$$\mathrm{B}\left[\%\right]=\left[1 - {\left({\displaystyle \frac{1000+{\mathsf{\delta }}^{13}\mathrm{C}}{1000+{\mathsf{\delta }}^{13}{\mathrm{C}}_0}}\right)}^{{\displaystyle \frac{1000}{\mathsf{\epsilon }}}}\right]·100$$

(1)

where δ¹³C and δ¹³C₀ are the carbon isotope values of the substrate at any given fraction of remaining substrate (C/C₀), and the initial substrate, respectively, and ε is the isotope enrichment factor. The δ¹³C value is defined as the deviation of the isotopic ratio of the chemical element E in the sample relative to the isotopic ratio of the reference standard, i.e., δ¹³C = R_sample/R_standard − 1 [31]. The enrichment factor (ε) represents the thousandth deviation of the isotopic fractionation factor (α) with respect to 1, i.e., ε = (α − 1) × 1000‰ [32]. Using the measured δ¹³C and δ¹³C₀ values, as well as the ε values obtained from the laboratory experiments, the degree of contaminant transformation (B [%]) can be calculated (Equation (1)) and used to assess the remediation performance of the contaminated site. This assessment is based on the isotopic enrichment factors, which could be variable due to the different environmental settings. This variability could introduce uncertainties into the evaluation of remediation effectiveness at contaminated sites. To effectively apply stable carbon isotope analysis in such assessments, it is crucial to comprehend the carbon isotope effects during the degradation process of pollutants and the impacts of the environmental conditions [33,34]. In addition, the isotopic enrichment factors can be a key tool for elucidating reaction mechanisms since different pathways of degradation could have distinct isotope effects, which can be reflected by their ε values [34,35,36].

In the study of photocatalytic degradation of TCE, although a large amount of work has been focused on catalyst optimization and degradation kinetic analysis [24,37,38], there is still a significant research gap regarding the systematic influence of environmental conditions on the carbon isotope enrichment factor (ε) and its relevance to the reaction mechanism. The existing literature mostly explores kinetic parameters or isotopic fractionation phenomena in isolation [24,34,38,39], with few studies combining the two to reveal the modulation of photocatalytic pathways by complex environmental factors (e.g., ionic species, redox conditions) [33,40]. For example, although it is known that NO₃^– and HCO₃^– can inhibit the degradation efficiency by quenching •OH, their quantitative effects on isotope fractionation and the consistency of the mechanism have not yet been clarified, and there is a lack of experimental validation of the stability of the ε values of different catalysts (e.g., TiO₂ and ZnO) under the same conditions to see if they reflect a universal reaction mechanism. This gap leads to elevated uncertainties in carbon isotope quantification models for field applications, limiting the reliability of photocatalytic techniques combined with compound–specific isotope analysis (CSIA). In this study, the influence of catalyst type, pollutant concentration, inorganic ions and oxygen concentration on the kinetic and isotope effects during the photodegradation of TCE was systematically investigated to fill the gaps in the theory of isotope fractionation of photocatalytic reaction, which provides an important scientific basis for the accurate remediation of complex polluted sites.

2. Materials and Methods

2.1. Materials

All chemicals are of analytical grade and were used as received: TCE (Macklin Biochemical Technology, Shanghai, China), TiO₂ (Aladdin Biochemical Technology, Shanghai, China, 99.8% metals basis, 5~10 nm, pure anatase, hydrophilicity), ZnO (Shandong Keyuan Biochemical, Shandong, China, 99%, 20~40 nm), and NaNO₃, NaHCO₃, Mg(NO₃)₂, and Ca(NO₃)₂ (Cologne Chemical, Chengdu, China). All aqueous solutions were prepared with deionized water.

2.2. Experiments of Photocatalytic Degradation of TCE

Photocatalytic reactions were conducted in a tin box (0.4 × 0.4 × 0.7 m) to maintain controlled environmental conditions. A water-cooling system stabilized the ambient temperature, ensuring experimental safety and reproducibility. A mercury lamp with wavelengths of 300~600 nm (500 W, model CEL-LAM500, CEC Jinyuan Science and Technology, Beijing, China) was placed at the center of the box. During the course of the experiment, the mercury lamp was cooled down by means of a water circulation unit. Each experiment was conducted in a 50 mL glass bottle, with a specific amount of TiO₂/ZnO added to ensure homogeneous dispersion in either 20 mL of deionized water or salt solutions of different concentrations. The glass bottles were sealed with a thick rubber stopper and flushed with helium (99.999%) for 10 min to purge the air from the headspace. TCE stock solution (1 g/L) was added using a syringe, and TCE was completely dissolved in water. Prior to irradiation, the suspension was magnetically stirred in the dark for 30 min. During irradiation, continuous magnetic stirring was maintained to prevent catalyst aggregation and sedimentation, and the reactor temperature was maintained stable using a water-cooling device. Due to its rapid equilibrium and very small equilibrium isotope fractionation factor [41], the concentration and isotope values of TCE vapor in headspace can represent those of TCE in solution. Headspace gas samples were collected with a gas-tight syringe at specific time intervals. The removal efficiency was calculated using Equation (2) [37].

$$\mathrm{Removal} \mathrm{efficiency}={\displaystyle \frac{{\mathrm{C}}_0-{\mathrm{C}}_{\mathrm{t}}}{{\mathrm{C}}_0}} \times 100\%$$

(2)

where C₀ (mg/L) and C_t (mg/L) represent the initial TCE concentration and TCE concentration at time t (h), respectively.

2.3. Analytical Methods

The carbon isotope composition and concentration of TCE were determined using a TRACE 1300 gas chromatograph (GC) coupled with an isotope ratio mass spectrometer (IRMS) MAT 253 (Thermo Fisher Scientific, Waltham, MA, USA). The GC was equipped with a capillary column HP-PLOT Q (Agilent technologies, Santa Clara, CA, USA, 30 m × 0.53 mm, 40 μm), and the GC inlet temperature was set at 200 °C. After the initial temperature of 40 °C was maintained for 2 min, the temperature was increased to 200 °C at a rate of 15 °C/min and maintained for 25 min. The carbon isotope values were reported relative to Vienna Pee Dee Belemnite (V-PDB). The TCE concentrations were calculated based on the peak areas of their IRMS signals. The error ranges of the experimental data were determined by three independent replicates: the standard deviation of δ¹³C was within ±0.2‰, and the relative error of concentration determination was within ±3%.

2.4. Calculation of Isotope Enrichment Factor (ε)

The extent and direction of isotopic fractionation are typically characterized by the isotope enrichment factor (ε). Based on the Rayleigh distillation model (a key tool for resolving mechanisms of isotope fractionation in the environment, linking changes in isotope ratios to changes in substance concentrations through mathematical relationships), it was obtained by plotting ln(C/C₀) versus ln[(1000 + δ¹³C)/(1000 + δ¹³C₀)] (Equation (3)) and determining the slope using least-squares regression [27]. The standard error is determined by the uncertainty of the slope, with values of ±0.1‰ to ±0.5‰.

$$\ln \left({\displaystyle \frac{1000+{\mathsf{\delta }}^{13}\mathrm{C}}{1000 +{\mathsf{\delta }}^{13}{\mathrm{C}}_0}}\right)={\displaystyle \frac{\mathsf{\epsilon }}{1000}} \times \ln {\displaystyle \frac{\mathrm{C}}{{\mathrm{C}}_0}}$$

(3)

where δ¹³C₀ and δ¹³C are the isotope values of initial TCE and residual TCE at time t (h), respectively. C₀ (mg/L) and C (mg/L) are concentrations of the initial TCE and residual TCE at time t (h), respectively.

3. Results and Discussion

3.1. Kinetic Effects of Environmental Conditions on TCE Photocatalytic Degradation

3.1.1. Catalyst Dosage and Types

To assess the effect of catalyst dosage on the photocatalytic degradation of TCE, different amounts of TiO₂ were added to solutions with 50 mg/L of TCE. As shown in Figure 1a, the degradation curves for all catalyst concentrations (0.5, 2.5, 5, and 25 g/L) revealed a rapid initial degradation of TCE, and slowed down afterwards. After 3 h, the degradation rates were 72%, 88%, 92%, and 82%, respectively. The degradation of TCE in all catalyst concentrations adhered to pseudo-first-order kinetics, R² > 0.97 (Figure 1b). At catalyst concentrations of 2.5 g/L and 5 g/L, the apparent rate constants were notably higher, with values of 0.84 h⁻¹ and 0.77 h⁻¹.

In order to investigate the effect of catalyst species on the photocatalytic degradation of TCE, the concentration of TCE solution (50 mg/L) was kept constant by adding different amounts of ZnO. The experimental results are shown in Figure 2a, where the degradation efficiencies of TCE were continuously increased with the increase in the catalyst concentration of ZnO (2.5 g/L, 5 g/L, and 25 g/L). The degradation efficiencies after 3 h were 87%, 89%, and 90%, respectively. When the concentration of ZnO is greater than 5 g/L, the degradation rate remains basically unchanged. The degradation of TCE followed pseudo-first-order kinetics (R² > 0.93) at all catalyst concentrations (Figure 2b). The maximum apparent rate constant of 0.92 h⁻¹ was observed when the initial concentration of the catalyst ZnO was 25 g/L.

Our results showed that the two catalysts, TiO₂ and ZnO, exhibited significant concentration-dependent differences for the photocatalytic degradation of TCE. At a catalyst concentration of 2.5 g/L, TiO₂ exhibited higher degradation efficiency with an apparent rate constant of 0.84 h⁻¹, which was better than that of 0.79 h⁻¹ for ZnO at the same concentration. This phenomenon was attributed to the better light absorption property (band gap width of 3.2 eV) and more efficient electron mobility of TiO₂ [42,43]. When the catalyst concentration was increased to 25 g/L, the degradation efficiency of ZnO was increased with an apparent rate constant of 0.92 h⁻¹, which exceeded that of 0.44 h⁻¹ for TiO₂. We suggest that the main reason for the decrease in TCE degradation efficiency at increasing TiO₂ catalyst dose (25 g/L) is the obstruction of UV light irradiation by the high concentration of TiO₂. At elevated TiO₂ doses, the cumulative presence of TiO₂ nanoparticles leads to substantial UV light scattering and absorption, thus preventing UV light from penetrating deeper into the solution [44]. This attenuation of UV light limits the extent of the photocatalytic reaction, which ultimately leads to a reduction in TCE degradation efficiency.

In photocatalytic systems, a moderate increase in photocatalysts can achieve higher photocatalytic activity due to the availability of ample active surface areas for TCE contact and participation in photochemical reactions. In addition, higher concentrations of electrons and holes contribute to the production of more •OH. However, an excess of catalyst concentration leads to adsorption aggregation, which diminishes the active surface areas significantly and impedes the catalyst’s absorption of UV light [20]. Therefore, it limits the generation of electron–hole pairs and reduces photocatalytic efficiency. Consequently, optimizing the catalyst dosage enhances the efficiency of photocatalytic degradation of TCE, ultimately reaching an optimal loading where the balance between light absorption and catalyst particle interactions is achieved [20,45].

3.1.2. TCE Concentrations

The impact of different TCE concentrations (25, 50, and 100 mg/L) on the photocatalytic degradation of TCE was investigated with TiO₂ concentration maintained at 2.5 g/L (0.05 g). Initially, TCE degradation was rapid, peaking at 0~1.5 h, after which the degradation rates slowed down (Figure 3a). After 2.5 h, the degradation rates were 86%, 90%, and 59% for initial TCE concentrations of 25, 50, and 100 mg/L, respectively. The initial degradation rates were 0.5 mg/(L·h) for all concentrations. The degradation of TCE at different concentrations followed the proposed first-order reaction kinetics, R² > 0.90 (Figure 3b). The apparent rate constant was highest at 50 mg/L (K_obs = 1.02 h⁻¹), and significantly slower at 100 mg/L (K_obs = 0.32 h⁻¹).

The initial concentration of TCE significantly influenced the photocatalytic degradation efficiency. An increase in the initial concentration of TCE from 25 mg/L to 50 mg/L resulted in a higher number of molecules exposed to active radicals, thereby enhancing the degradation rate initially. However, when the initial concentration of TCE was further increased to 100 mg/L, the degradation rate decreased.

3.1.3. O₂

The presence of O₂ in the photocatalytic reaction may also contribute to photocatalytic degradation. In order to investigate the effect of O₂ on the degradation of TCE, we conducted the following experiments: (1) blank control experiments were performed under dark conditions, where only O₂ (with an estimated molar ratio of 5:1 to TCE) was supplied; (2) a 2.5 g/L TiO₂ catalyst was introduced with UV irradiation with only O₂ in the headspace (estimated 5:1 molar ratio to TCE). The results were compared with those of a control experiment without O₂, in which helium of 99.999% purity was continuously fed into the reactor for 10 min to remove O₂ in the headspace and most of the dissolved oxygen. As shown in Figure 4, under dark conditions, even in the presence of sufficient O₂, no degradation of TCE was detected within 2 h. In contrast, with the addition of O₂, the catalyst, and the presence of UV illumination, the degradation efficiency of TCE was 95% within 2 h, which was significantly faster than that of the case with no O₂ exposure (89%). The accelerated degradation efficiency can be attributed to the capture of photogenerated electrons by O₂ as a highly efficient electron acceptor to generate superoxide radicals (O₂•⁻), and the simultaneous generation of reactive oxygen species (Equations (4)–(6)) such as hydrogen peroxide (H₂O₂) and peroxyhydroxyl radicals (HO₂•) through chain reactions [23,46]. These strong oxidizing species, besides •OH, could also degrade TCE. It is also possible that O₂ is involved in the subsequent reactions of TCE degradation, which in turn accelerates the degradation efficiency.

$${\mathrm{O}}_2+{\mathrm{e}}^{-}\to {\mathrm{O}}_2{•}^{-}$$

(4)

$${\mathrm{O}}_2{•}^{-}+{ \mathrm{H}}^{+}\to {\mathrm{HO}}_2•$$

(5)

$${\mathrm{O}}_2{•}^{-}+{\mathrm{HO}}_2•+{ \mathrm{H}}^{+}\to {\mathrm{H}}_2{\mathrm{O}}_2+{\mathrm{O}}_2$$

(6)

3.1.4. Concentrations of Anions

Inorganic anions such as NO₃⁻ and HCO₃⁻ are widely present in natural waters, and due to geographic and anthropogenic influences, their concentrations may vary from 10⁻⁵ to 10⁻³ mol/L [47,48]. However, in some polluted waters, the concentrations of NO₃⁻ and HCO₃⁻ could be even higher [49,50]. In the photocatalytic reaction, inorganic anions could react with the active substance, affecting its concentration and properties [51,52], leading to changes in TCE removal efficiency. To investigate the effects of inorganic anions on the photocatalytic degradation of TCE, the TCE concentration was kept constant at 50 mg/L, and the concentration of TiO₂ was maintained at 2.5 g/L, while anions of different concentrations were added for comparison experiments.

• NO₃⁻ Concentrations

The experimental results are shown in Figure 5a, where the removal of TCE was 84% (0.002 mol/L NO₃⁻), 83% (0.02 mol/L NO₃⁻), and 76% (0.2 mol/L NO₃⁻), respectively, after 2.5 h. The photocatalytic degradation of TCE was more obviously inhibited when the NO₃^– concentration was 0.2 mol/L, and less inhibited when the NO₃^– concentration was 0.002–0.02 mol/L.

In photocatalytic reactions, UV light can activate NO₃⁻ to produce •OH (Equations (7) and (8)) [52], which increases the •OH concentration in the reaction and enhances the efficiency of photocatalytic degradation of TCE. However, it was found that NO₃⁻ hindered the photodegradation of TCE during the experiments. This may be because NO₃⁻ also consumes •OH to produce NO₃• (Equation (9)) [53]. The reduction potential of NO₃• (2.3~2.4 V) is lower than that of •OH (2.7 V) [54], which produces a certain inhibitory effect. Meanwhile, with the increase in NO₃⁻ concentration in the solution, the concentration of NO₃⁻ adsorbed on the catalyst surface by electrostatic action will also increase, competing with TCE molecules for the active site, reducing the contact area between TCE and the catalyst, and thus decreasing the degradation efficiency [55].

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

(7)

$${\mathrm{O}•}^{-}+{\mathrm{H}}_2\mathrm{O}\to •\mathrm{OH} + {\mathrm{OH}}^{-}\hspace{1em}\hspace{1em}\mathrm{k}=1.8 \times { 10}^6 {\mathrm{M}}^{-1}{\mathrm{s}}^{-1}$$

(8)

$${\mathrm{NO}}_3^{-}+•\mathrm{OH}\to {\mathrm{NO}}_3•+{\mathrm{OH}}^{-}$$

(9)

• HCO₃⁻ concentrations

Figure 5b indicates that the variations in HCO₃⁻ concentrations exerted an inhibitory effect on the photocatalytic degradation of TCE. Within 2.5 h, the degradation efficiencies of TCE were 84% at 0.002 mol/L, 71% at 0.02 mol/L, and 58% at 0.2 mol/L.

An increase in the concentration of HCO₃⁻ scavenges •OH and produces CO₃•⁻ (Equation (10)) [46]. The reduction potential of CO₃•⁻ (1.8 V) is lower than that of •OH (2.7 V) [56], which produces a certain inhibitory effect. As the concentration of HCO₃⁻ increases, the concentration adsorbed on the surface of the catalyst increases, reducing the contact area between the TCE and the catalyst, and thus the degradation efficiency [55]. Meanwhile, a high concentration of HCO₃⁻ will increase the solution pH, which affects the surface charge of the catalyst [57], thus affecting its adsorption capacity and electron transfer capacity, and further affecting the catalytic activity of the catalyst. It has been shown that the solution alkalinity caused by carbonate and bicarbonate can lead to the aggregation of TiO₂, and the particle size increases significantly from 3 μm to 450 μm, which reduces the catalytic activity of TiO₂ [58].

$${\mathrm{HCO}}_3^{- }+•\mathrm{OH}\to {\mathrm{CO}}_3{•}^{-}+{\mathrm{OH}}^{-}\hspace{1em}\mathrm{k}=3.5 \times { 10}^5{ \mathrm{M}}^{-1}{\mathrm{s}}^{-1}$$

(10)

3.1.5. Cation Types

Cations such as Mg²⁺ and Ca²⁺ are commonly found in natural water bodies. To explore the effect of cationic species on the photocatalytic degradation of TCE, experiments were conducted under constant conditions, with the addition of 0.02 mol/L of Mg²⁺ or Ca²⁺ at 50 mg/L TCE. As shown in Figure 6a, the degradation efficiencies were 89% (0.02 mol/L Mg²⁺) and 87% (0.02 mol/L Ca²⁺) after 2.5 h. The degradation of TCE followed the proposed first-order reaction kinetics, with R² > 0.93 (Figure 6b). The apparent rate constants were highest in the absence of cations (K_obs = 1.02 h⁻¹) and were 0.79 h⁻¹ and 0.74 h⁻¹ in the presence of 0.02 mol/L Mg²⁺ or Ca²⁺, respectively.

The results suggest that there exists a small inhibitory effect on the photocatalytic degradation of TCE in the presence of cations. This is because Mg²⁺ and Ca²⁺ present in the solution will adsorb on the surface of the catalyst by electrostatic action, competing with the TCE molecules for the active sites, reducing the contact area between the TCE and the catalyst, and thus reducing the degradation efficiency [55].

3.2. Isotope Effects of Environmental Conditions on Photocatalytic Degradation of TCE

A series of experiments was conducted to investigate the influence of environmental conditions on the carbon isotope effects during the photocatalytic degradation of TCE. Catalyst dosage and types, concentrations of TCE, O₂, and ions (e.g., NO₃⁻, HCO₃⁻, Mg²⁺, and Ca²⁺) were selected as variables.

3.2.1. Catalyst Dosage and Types

As shown in Figure 7a, the ε values change very little with different concentrations of the catalyst, within the range of −2.3 ± 0.3‰ to −3.2 ± 0.5‰. Different TiO₂ dosages affect the production of •OH, thus the overall degradation efficiency of TCE. However, the photocatalytic process of generating •OH for the oxidative degradation of TCE remains unchanged. Given that isotope effects mainly depend on the degradation mechanism, the small change in ε values suggests that changes in catalyst dosage did not significantly impact the degradation mechanism.

By changing the types of catalyst, replacing TiO₂ with ZnO, the ε values of the system showed significant stability characteristics. As shown in Figure 7b, when the ZnO concentrations were varied in the range of 2.5~25 g/L, the ε values only varied in the range of −2.0 ± 0.1‰ to −2.3 ± 0.4‰, and the ε values of each concentration condition were close to those of the TiO₂ system (−2.1 ± 0.2‰ to −3.2 ± 0.3‰), which indicated that the change of catalyst type did not lead to the change of degradation mechanism.

The ε values of TCE photocatalytic degradation in this study are comparable to those of the degradation of TCE by a Fenton reaction (ε = −2.8‰ to −3.3‰) [34,59] and Fenton-like reaction (ε = −2.7‰ to −3.6‰) [60]. Kinetic isotope effects are primarily determined by the mechanism of chemical bond breaking [27,51,60]. Photocatalytic degradation of TCE shares a similar mechanism with TCE degradation by a Fenton reaction and Fenton-like reaction. These reactions all involve the oxidation of TCE in the form of •OH attacking C=C bonds [61,62]. However, TCE degradation by potassium permanganate oxidation (ε = −17‰ to −26.8‰) [63,64], zero-valent iron reduction (ε = −8.6‰ to −14.8‰) [65,66,67,68], and certain microbial reductive dechlorination (ε = −4.07‰ to −16.4‰) exhibited significantly larger isotope effects. The oxidation of TCE by potassium permanganate is also an addition reaction, but the C=C reacts with MnO₄⁻ to form a cyclic hypomanganate instead [69]. In contrast, microbial reduction and other abiotic reductive dechlorination, which occurs through the breaking of carbon and chloride bonds as an initial reaction step, differ significantly from photocatalytic reactions in terms of reaction mechanism, resulting in different ε values [70]. Carbon isotope enrichment factors (ε = −2.4‰ to −4.2‰ and −2.4 to −3.4‰) have been reported for TCE degradation by anaerobic dechlorination [71,72], which are similar to those of photocatalytic degradation (ε = −2.0‰ to −3.2‰), despite the different degradation mechanisms. This similarity may be attributed to the fact that different mechanisms could produce similar carbon isotope effects values. It is also possible that, because of the different levels of commitment to catalysis in these degradation reactions, the variability in isotopic fractionation associated with the cleavage of different bonds may be cancelled out [57] (

Table 1. Isotopic enrichment factor ε for abiotic and biotic oxidation of TCE.

Reaction Type	ε (‰)	Reference
Photocatalytic	−2.0 to −3.2	This study
Fenton	−2.8 to −3.3	[34,59]
Fenton-like	−2.7 to −3.6	[60]
Permanganate oxidation	−17 to −26.8	[63,64]
Anaerobic reduction by certain microorganisms	−2.4 to −4.2	[71,72]

).

3.2.2. TCE Concentrations

In order to investigate the effect of TCE concentration changes on carbon isotope fractionation, three sets of experiments with different TCE concentrations (25 mg/L, 50 mg/L, 100 mg/L) were set up in this study. The results are shown in Figure 8. With the different TCE concentrations, the ε values varied very little, ranging from −2.6 ± 0.2‰ to −3.6 ± 0.3‰, and the ε value under each concentration condition was close to those of the catalyst system (−2.0 ± 0.1‰ to −3.2 ± 0.3‰), which indicated that the change of TCE concentration did not result in the change of the degradation mechanism.

3.2.3. O₂

We investigated the isotope effects during the photocatalytic degradation of TCE with the presentation of O₂ (molar ratio of TCE to O₂ estimated to be 1:5), and the ε value for TCE degradation was measured to be −2.8 ± 0.1‰ (Figure 9), which is similar to the enrichment factor under anaerobic conditions (ε = −3.2 ± 0.5‰). Although the introduction of O₂ promoted the generation of reactive oxygen species and increased the degradation efficiency to 95%, the isotopic enrichment factor did not change significantly. This may be attributed to the fact that even though O₂ was involved in the generation of secondary oxidants such as O₂•⁻ and H₂O₂, the dominant role of •OH was not altered, and the degradation pathway was still dominated by •OH. The stability of this isotopic enrichment factor provides strong evidence for the reaction mechanism of photocatalytic degradation of TCE, suggesting that O₂ mainly enhances the degradation efficiency without changing the pathway of the reaction.

3.2.4. Concentrations of Anions

• NO₃⁻ Concentrations

Comparative experiments were conducted by introducing solutions of NO₃⁻ at concentrations of 0.002 mol/L, 0.02 mol/L, and 0.2 mol/L. The results are shown in Figure 10a with ε values of −2.4 ± 0.2‰, −3.4 ± 0.5‰, and −2.6 ± 0.1‰, respectively. Compared to the isotope effects in deionized water (ε = −3.2 ± 0.5‰), the differences were not significant. These results suggested that changes in NO₃⁻ concentrations have small effects on the carbon isotope effects of TCE photocatalytic degradation. The photocatalytic degradation of TCE is affected by different concentrations of NO₃⁻, which scavenge •OH. However, changes in these concentrations do not affect the mechanism of TCE photocatalytic degradation, probably because it degrades TCE with ε values similar to that of the •OH, and/or the concentrations of NO₃• produced are relatively low, so they do not play an important role in the degradation. Similar observations were made by Liu et al. [34] in experiments on the Fenton-like degradation of TCE, where only minor differences in ε values (−3.3 ± 0.4‰) were observed at different initial reaction conditions (6 mol/L NO₃⁻).

• HCO₃⁻ Concentrations

In order to investigate the effect of HCO₃⁻ on carbon isotope fractionation during the photocatalytic reaction for the degradation of TCE, an experimental comparative analysis was carried out to determine the carbon isotope effects at different HCO₃⁻ concentrations. The results are shown in Figure 10b, and the ε values corresponding to 0.002 mol/L, 0.02 mol/L, and 0.2 mol/L of HCO₃⁻ were −3.2 ± 0.4‰, −3.4 ± 0.4‰, and −4.4 ± 0.3‰, respectively. It can be observed that when the HCO₃⁻ concentrations (0.002 to 0.02 mol/L) were low, the ε values were relatively stable and basically unchanged, but when the HCO₃⁻ concentration (0.2 mol/L) increased, the ε value showed a small change. This may be due to the fact that one of the final products of TCE degradation is CO₂, which dissolves in water to produce HCO₃⁻. Excess HCO₃⁻ in the solution may have some effects on our whole TCE photocatalytic reaction process, leading to a small deviation in the isotopic enrichment factor. Similar results were observed by Liu et al. [33] in the degradation of TCE with zero-valent iron activated persulfate, with only minor differences in the ε values (−3.4 ± 0.3‰ to −4.3 ± 0.3‰) at different initial reaction conditions (10 mmol/L and 100 mmol/L HCO₃⁻).

3.2.5. Cation Types

In this experiment, the effects of cations (Mg²⁺ and Ca²⁺) on the carbon isotope effect of photodegradation of TCE were investigated. The results are shown in Figure 11; the ε values were determined to be −2.1 ± 0.3‰ for Mg²⁺ and −2.8 ± 0.2‰ for Ca²⁺, which were not significantly different from that of deionized water (−3.2 ± 0.5‰).

The results show that the ε values of adding Mg²⁺ and Ca²⁺ are not significantly different from those of deionized water, and they do not significantly affect the carbon isotope effects in the photocatalytic degradation of TCE, and do not change the mechanism of TCE degradation.

4. Conclusions

In this study, the effects of different conditions on the kinetics and carbon isotope effects of photocatalytic degradation of TCE were systematically investigated. Increasing the concentration of TCE, anions (NO₃⁻ and HCO₃⁻), and cations (Mg²⁺ and Ca²⁺) decreased the photocatalytic degradation of TCE. O₂ increased the degradation efficiency of TCE. The dose required to achieve maximum photocatalytic efficiency varies for different types of catalysts, which needs to be determined on a case-by-case basis. The carbon isotope enrichment factor for photocatalytic degradation by TCE was determined for the first time, with ε values ranging from −2.0 ± 0.2‰ to −3.2 ± 0.5‰. Concentrations of catalysts (TiO₂ and ZnO), TCE, O₂, nitrate, bicarbonate, and metal ion species (Mg²⁺ and Ca²⁺) had little to small effects on the ε values. The ε values are stable and reproducible, greatly reducing the uncertainty associated with the application of isotope enrichment factors to assess the effectiveness of remediation by photocatalytic degradation reactions. Future work could further investigate the influence of other environmental factors (e.g., pH, other inorganic ions, different light sources, light intensity, etc.) on the reaction pathways and isotope effects of TCE photocatalytic degradation.