Stable Carbon Isotope Fractionation of Trichloroethylene Oxidized by Potassium Permanganate Under Different Environmental Conditions

Abstract

Stable isotope analysis is a powerful tool for inferring and quantifying transformation processes, but its effectiveness relies on understanding the magnitude and variability of isotopic fractionation associated with specific reactions. Potassium permanganate (KMnO₄) is widely used as an efficient oxidant for the degradation of trichloroethylene (TCE); however, the influence of environmental factors on the isotope fractionation during this process remains unclear. In this study, compound-specific isotope analysis (CSIA) was conducted to investigate the variability in carbon isotope effects during the KMnO₄-mediated degradation of TCE under varying conditions, including initial concentrations of KMnO₄ and TCE, the presence of humic acid (HA), pH levels, and inorganic ions. The results showed that the overall carbon isotope enrichment factors (ε) of TCE ranged from −26.5 ± 0.5‰ to −22.8 ± 0.9‰, indicating relatively small variations across conditions. At low KMnO₄/TCE molar ratio (n(KMnO₄)/n(TCE)), incomplete oxidation and/or MnO₂-mediated oxidation of TCE likely resulted in smaller ε. For dense, non-aqueous phase liquid (DNAPL) TCE, which represents extremely high concentrations, the ε value was −13.0 ± 1.7‰ during KMnO₄ oxidation. This may be attributed to the slow dissolution of isotopically light TCE from the DNAPL phase, altering the δ¹³C signature of the reacted TCE and resulting in a significantly larger ε value than observed for dissolved-phase TCE oxidation. The ε values increased with rising pH, probably due to the decrease in oxidation potential (E₀) of KMnO₄ from pH ~2 to ~12, as well as the emergence of different degradation pathways and intermediates under varying pH conditions. Both SO₄²⁻ and NO₃⁻ slightly influenced the ε values, potentially due to the formation of H₂SO₄ and HNO₃ at lower pH, which may act as auxiliary oxidants and contribute to TCE degradation. A high concentration (50 mM) of HA led to a decrease in ε values, likely due to competitive interactions between HA and TCE for KMnO₄, which reduced the effective oxidation of TCE. Overall, the carbon isotope enrichment factors for KMnO₄-mediated TCE degradation are relatively stable, although certain environmental conditions can exert minor influences. These findings highlight the need for caution when applying quantitative assessment based on CSIA for KMnO₄ oxidation of TCE.

1. Introduction

Since trichloroethylene (TCE) is one of the most ubiquitous and carcinogenic pollutants in contaminated groundwater [1], its migration from subsurface environments into the indoor air of overlying buildings is one of serious concern [2]. TCE is chemically stable and resistant to biodegradation in aquifers due to its low viscosity, high density, low solubility, relatively high volatility and strong affinity for subsurface materials [3], making it a persistent, long-term pollution source of contamination. TCE can penetrate below the water table as a dense, non-aqueous phase liquid (DNAPL), from which it may slowly dissolve over decades [4,5], posing sustained threats to groundwater quality [6,7,8]. Consequently, remediation efforts often require extended timescales, sometimes spanning decades [1]. Various remediation strategies have been explored at a laboratory scale to remove TCE from contaminated soil and groundwater. Environmentally friendly approaches, such as in situ biodegradation and phytoremediation, driven by natural or solar processes, are considered ideal [9,10,11]. However, high concentrations of contaminants can inhibit microbial activity, and many bioremediation methods require significantly longer timeframes compared to chemical treatments [9,10,11]. Pump-and-treat technologies often fail to achieve complete contaminant removal and necessitate prolonged site management due to contaminant rebound and tailing effects [12,13]. In contrast, chemical oxidation has gained widespread application for groundwater and aquifer remediation over the past several decades [14,15,16,17]. This method can degrade a broad spectrum of organic pollutants via electron transfer or generation of reactive free radicals (mainly including •OH, SO₄•−, O₂•−, and HO₂•), which have strong oxidizing properties [18,19,20,21,22].

Among the various oxidants, including Fenton’s reagent, ozone, potassium permanganate (KMnO₄) and persulfate [23], KMnO₄ stands out for its strong oxidative potential, high efficiency, low cost, aqueous stability, broad applicability, and formation of nontoxic byproducts [3,17,24,25,26]. KMnO₄ has been widely used in drinking water and wastewater treatment for decades. By the late 20th century, it has also become a popular choice for in situ chemical oxidation (ISCO) [27,28], offering a rapid and cost-effective treatment strategy for zones of high residual contamination. A survey of ISCO applications reported that KMnO₄ was widely used in remediation projects [29]. It has proven effective in degrading a wide variety of organic compounds, including chlorinated solvents, chlorophene, triclosan and bromophenols, across a broad pH range [30,31,32,33,34]. Permanganate ions (MnO₄⁻) are particularly suited for oxidizing organics contaminants that contain carbon–carbon double and single bonds, aldehyde, or hydroxyl functional groups, including chlorinated ethenes [30,35]. Compared with Fenton’s reagent, KMnO₄ often demonstrates higher removal efficiency for some chlorinated solvents [36]. Its stability and durability allow MnO₄⁻ to remain reactive and mobile, facilitating contact with subsurface contaminants [37,38]. KMnO₄ acts as a strong oxidant, rapidly degrading chlorinated solvents such as TCE and perchloroethylene (PCE) into MnO₂, CO₂, and Cl⁻ as the primary reaction products [39,40]:

C₂HCl₃ + 2MnO₄⁻ → 2MnO₂(s) + 2CO₂ + 3Cl⁻ + H⁺

(1)

The widely accepted mechanism for the oxidation of TCE by KMnO₄ involves an electrophilic attack on the carbon-carbon double bond, forming a cyclic hypomanganate diester intermediate [41]. For TCE, kinetic modeling indicates that the formation of the intermediate is the rate-limiting step in the overall reaction [42]. The identity of the oxidation product is highly dependent on pH, with proposed reaction pathways including hydrolytic decomposition of the cyclic ester or direct cleavage by MnO₄⁻ to produce two formic acid molecules [41,43]. Under certain conditions, all carboxylic intermediates may undergo further oxidation to carbon dioxide (CO₂) [16,39,44].

In recent years, stable isotope fractionation has gained increasing attention as a tool to determine the magnitude and variability of transformation resulting from different degradation processes and environmental conditions [40,42,45,46]. Compound-specific isotope analysis (CSIA) has emerged as a powerful analytical tool for evaluating the effectiveness of in situ remediation of organic contaminants at contaminated sites [29,45,47,48]. This isotope analysis is commonly based on the Rayleigh model (Equation (2)), which describes the relationship between isotope fractionation and changes in contaminant concentration due to specific transformation mechanisms [48]. By applying values for δ¹³C₀ (initial isotope composition) and δ¹³C (remaining substrate isotope composition after reaction), along with the carbon isotope enrichment factor (ε) derived from laboratory experiments, the extent of contaminant transformation (B[%]) can be calculated using Equation (3). This enables a quantitative assessment of remediation performance [49,50].

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

(2)

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

(3)

where δ¹³C is the isotope ratio of the TCE at any given fraction of TCE remaining (ƒ), and δ¹³C₀ is the initial value.

Previous studies have documented carbon isotope fractionation during permanganate oxidation of TCE under varying concentrations of chloride, KMnO₄, and TCE [40]. However, the influence of environmental factors on isotope fractionation during this process remains unclear. The objective of this study was to evaluate the effect of environmental variables on carbon isotope fractionation during KMnO₄-mediated degradation of TCE. Specifically, we examined the impact of varying KMnO₄ and TCE concentrations, the presence of HA, a broad pH range (2–12), and several common inorganic ions (NO₃⁻, HCO₃⁻, SO₄²⁻, and Cl⁻), which collectively represent typical environmental conditions in natural waters. HA may play a competitive role with TCE oxidized by KMnO₄ in this study due to its complex structure (high-molecular-weight subunits, carbon−carbon double bonds, π-electrons, and long conjugated bonds) [51,52] Neglecting the influence of these environmental factors could introduce uncertainty in the application of stable carbon isotope analysis for evaluating ISCO performance. The findings of this study aim to support the reliable use of stable carbon isotope analysis as a quantitative tool for assessing ISCO-based remediation efforts. This study systematically explores the stable carbon isotope fractionation of TCE oxidized by KMnO₄ under different environmental conditions.

2. Materials and Methods

2.1. Materials

Trichloroethylene (TCE, 99.9%), humic acid (HA, ≥90%), and potassium permanganate (KMnO₄, AR grade) were purchased from Shanghai Macklin Biochemical Technology Co., Ltd., (Shanghai, China). Sodium hydroxide (NaOH, ≥96.0%) was obtained from Lianlong Bohua (Tianjin, China) Pharmaceutical & Chemical Co. (Tianjin, China). Other reagents, including Sodium chloride (NaCl, ≥99.5%), sodium nitrate (NaNO₃, ≥99.5%), disodium hydrogen phosphate dodecahydrate (Na₂HPO₄·12H₂O, ≥99.0%), sodium dihydrogen phosphate dihydrate (NaH₂PO₄·2H₂O, ≥99.0%), and sodium phosphate dodecahydrate (Na₃PO₄·12H₂O, ≥98.0%) were purchased from Chengdu Cologne Chemical Co., (Chengdu, China). Phosphoric acid (H₃PO₄, ≥85.0%) was provided by Tianjin Kaitong Chemical Reagent Co. (Tianjin, China). Sodium bicarbonate (NaHCO₃, AR grade) was sourced from Sichuan Xilong Scientific Company (Chengdu, China). All chemicals were used as received without further purification.

2.2. Experimental Procedure

A series of batch experiments were conducted at room temperature to evaluate the variability of carbon isotope fractionation during KMnO₄-mediated TCE degradation. KMnO₄ stock solution (10 g/L) was prepared by dissolving 0.5 g KMnO₄ crystals in 50 mL deionized water in a volumetric flask. A near-saturated TCE stock solution (1000 mg/L) was prepared by adding 340 μL of pure-phase TCE to 500 mL of deionized water in a 500-mL anaerobic flask, followed by in the oscillator at 25 °C 180 rpm for 12 h. To initiate the reaction, 18 mL phosphate buffer solution (pH ~8.0) was placed in a 65 mL anaerobic flask and sealed with a PTFE membrane rubber gasket. Then, 1.5 mL of TCE stock solution (1000 mg/L) was injected into each vial. Subsequently, 19.5 mL of the reaction mixture was added to 65 mL conical flasks, and kinetic reactions were initiated by adding 0.5 mL of KMnO₄ stock solution (10 g/L), resulting in a final TCE concentration of to the 75 mg/L. Kinetic experiments were conducted under ambient temperature with stirring at 500 rpm. Prior to KMnO₄ addition, the mixture was allowed to equilibrate for 1 h to establish gas-liquid equilibrium and the headspace gas was sampled to determine the initial TCE concentration. In the batch experiments, the effects of various parameters were investigated, including pH (2–12, adjusted using phosphate buffer), KMnO₄ dose (50–400 mg/L), TCE concentration (25–250 mg/L), inorganic ions (0.2 M, including Cl⁻, HCO₃⁻, NO₃⁻, and SO₄²⁻), and HA (0.5–250 mM). HA stock solution (1 M) was prepared by mixing 11.35 g of HA (sieved through 200-mesh to ensure uniform particle size) with 50 mL of deionized water and in the oscillator at 25 °C 180 rpm for approximately 24 h.

2.3. Analysis Methods

Due to the rapid gas–liquid equilibrium of TCE and the minimal isotope fractionation associated with phase transfer, the headspace TCE concentration was considered representative of the aqueous phase concentration [53,54]. At each sampling point, 500 μL of headspace gas was collected using a gastight glass syringe. The carbon isotope composition of TCE was analyzed using a TRACE 1300 gas chromatograph (GC) coupled with a MAT 253 isotope ratio mass spectrometer (IRMS). The GC was equipped with a DB-624 capillary column (30 m × 0.25 mm, 1.40 μm film thickness), and the inlet temperature was set at 200 °C. The oven temperature was initially held at 60 °C for 1.5 min, ramped to 200 °C at 15 °C/min, and held at 200 °C for 25 min. The oxidation reactor temperature was maintained at 960 °C. Carbon isotope values were reported relative to the Vienna Pee Dee Belemnite (V-PDB) standard. TCE concentrations were quantified based on peak areas of their chromatographic IRMS signals.

3. Results and Discussion

3.1. Initial Concentration of TCE and KMnO₄

Figure 1a,d showed that TCE removal efficiencies strongly depended on the initial concentrations of KMnO₄ and TCE. The TCE oxidation by KMnO₄ was monitored until its concentrations fell below the detection limit. As shown in Figure 1a, the half-lives of TCE degradation at KMnO₄ concentrations of 50, 150, 250 and 400 mg/L were approximately 120, 21, 15, and 12 min, respectively. Higher KMnO₄ concentrations led to faster degradation rates and higher final removal efficiencies, consistent with Zhang and Dong [44]. Figure 1d indicates that after 60 min, the removal efficiencies for initial TCE concentrations of 25, 75, 150, and 250 mg/L were 92%, 90%, 83%, and 74%, respectively. These results confirm that TCE removal is highly dependent on both the initial concentrations of KMnO₄ and TCE. In other words, the molar ratio of KMnO₄/TCE can influence the removal efficiency. When 75 mg/L TCE was treated with 50–400 mg/L KMnO₄, the molar ratios of KMnO₄ to TCE (n(KMnO₄)/n(TCE)) ranged from 0.55–4.42. When 250 mg/L KMnO₄ was used with 25–250 mg/L TCE, the molar ratios ranged from 0.83–8.29. According to proposed mechanisms, the initial step of TCE oxidation involves cleavage of the C=C bond, forming an acyclic ester and consuming 1 mol of KMnO₄ per mol of TCE [41]. However, complete mineralization into CO₂ requires further oxidation of intermediate carboxylic acids (formic, oxalic and glyoxylic acids), consuming an additional 1 mol of KMnO₄ [16]. Thus, when n(KMnO₄)/n(TCE) exceeds 1, complete TCE removal can be achievable. This was observed for 75 mg/L TCE with KMnO₄ concentrations of 150–400 mg/L. However, with only 50 mg/L KMnO₄ (n(KMnO₄)/n(TCE) = 0.55), the removal efficiency dropped to around 50%, even though the molar ratio is close to the theoretical 1:1 for the first oxidation step. Interestingly, 250 mg/L (1.58 mM) KMnO₄ was able to degrade 250 mg/L (1.91 mM) TCE with 88.5% efficiency, despite a sub-stoichiometric molar ratio of 0.83. This suggests additional degradation pathways may be at play, such as the involvement of MnO₂, a byproduct of KMnO_4, which can further oxidize TCE [55]. The decreasing reaction rate at higher TCE concentrations may result from increased MnO₂ colloid formation, which contributes to oxidation but may also limit accessibility or slow reaction dynamics.

Although high KMnO₄ concentrations improved the degradation rate, their effect on carbon isotope fractionation was investigated through the enrichment factor ε values. As shown in Figure 1b,e, δ¹³C values of TCE changed with increasing KMnO₄ and TCE concentrations. Figure 1c shows ε values increasing from −25.2 ± 1.3‰ to −22.8 ± 0.9‰ as KMnO₄ concentration increased from 50–400 mg/L. In contrast, Figure 1f shows a slight decreasing trend in ε value from −22.9 ± 1.1‰ to −26.5 ± 0.5‰ as initial TCE concentrations increased from 25–250 mg/L. These values align with previously reported ranges where ε values were found between −26.3 ± 2.4‰ and −20.9 ± 1.1‰ [40], while Hunkeler, Aravena, Parker, Cherry and Diao [42] reported −25.1 ± 0.4‰ for excess KMnO₄ (1250 mg/L) and −26.8 ± 1.1‰ under limited KMnO₄ (67–157 mg/L) supply.

While previous studies concluded that ε values remain unaffected by excess or insufficient TCE, our results suggest a subtle dependence on the n(KMnO₄)/n(TCE) ratio (Figure 2). As this ratio increases, ε values also tend to increase, possibly due to the shifts in relative contributions of KMnO₄ and MnO₂ in the oxidation process. Under KMnO₄-limited conditions (n(KMnO₄)/n(TCE) < 2), MnO₂ colloids formed during the reaction may oxidize residual TCE, slightly altering the carbon isotope effect [41,55]. When TCE was excess, KMnO₄ was quickly consumed by TCE, and the proportion of MnO₂ oxidizing TCE was elevated. There may be a difference between the ε values of KMnO₄ oxidized TCE and MnO₂ oxidized TCE, resulting in a small change in the overall ε values at a higher degree of participation of MnO₂ oxidized TCE, which needs to be further investigated.

Another potential explanation is that insufficient KMnO₄ leads to incomplete oxidation, which could affect ε values. Although earlier studies overlooked this, Poulson and Naraoka [40] observed a lower ε value (−26.3 ± 2.4‰) in a system with incomplete oxidation compared to one with sufficient oxidant, supporting our observations.

To further evaluate these effects under extreme TCE loading, we examined the isotope behavior of DNAPL-phase TCE and a near-saturated aqueous TCE solution (950 mg/L). Compared to 75 mg/L TCE degraded by 250 mg/L KMnO₄, TCE degradation was significantly lower in the DNAPL and 950 mg/L conditions, even when using 500 mg/L KMnO₄ (Figure 3a). Furthermore, Figure 3b shows a corresponding decrease in δ¹³C values with increasing TCE concentrations. Figure 3c shows that ε = −24.4 ± 0.9‰ for 75 mg/L TCE and −23.4 ± 2.1‰ for 950 mg/L TCE. However, the ε value for DNAPL TCE was significantly higher at −13.0 ± 1.7‰. This discrepancy likely results from the slow dissolution of the DNAPL phase, which skews the δ¹³C value of the reacting TCE pool. As heavier TCE accumulates in the aqueous phase and lighter isotopes are continuously supplied via dissolution, the headspace δ¹³C deviates from the Rayleigh model, producing an apparent increase in ε values.

The evolution of TCE degradation of DNAPL indicated that the TCE is obviously overdosed, but the TCE concentration of headspace gradually decreased for a long time. This may be because the rate of liquid TCE dissolved into the high concentration of TCE solution was slow. The concentration of TCE is still high even if the KMnO₄ oxidized a part of the TCE. Compared with the degradation curve of 950 mg/L TCE, the same concentration of KMnO₄ can only react with about half of the saturated solution of TCE. The saturated solution could not be reached within the time of the experiment.

These findings suggest that in a fully dissolved system, increasing TCE concentrations have a small impact on the isotope effect. However, in DNAPL-contaminated zones where dissolution is rate-limiting, isotope effects are more pronounced and may serve as indicators of source zone contamination where TCE persists in the non-aqueous phase.

3.2. Initial pH

In general, the pH of the solution significantly affects the reaction kinetics of KMnO₄-mediated TCE degradation. This is primarily due to the decreasing oxidation potential (E₀) of KMnO₄ with increasing pH [17,56], which leads to a slower TCE oxidation rate.

MnO₄⁻ + 5e⁻ + 8H⁺ = Mn²⁺ + 4H₂O E₀ = 1.51 V (pH < 3.5)

(4)

MnO₄⁻ + 3e⁻ + 2H₂O = MnO₂ + 4OH⁻ E₀ = 0.59 V (3.5 < pH < 12)

(5)

MnO₄⁻ + e⁻ = MnO₄²⁻ E₀ = 0.56 V (pH > 12)

(6)

Figure 4a shows that after a 110 min reaction, the TCE degradation efficiencies were similar across the pH range of 2 to 12. This may be attributed to the excess KMnO₄, which is capable of oxidizing nearly all TCE present [17]. However, the oxidation rate was lowest at pH ~12, likely due to the reduced E₀ of KMnO₄ at a high pH. Additionally, the rate constants for the oxidation of TCE intermediates are known to strongly depend on pH, particularly in the range of 4–8 [41]. These intermediates are oxidized more rapidly at lower pH, facilitating faster CO₂ formation.

As shown in Figure 4b, the ε values of KMnO₄-mediated TCE degradation ranged from −26.5 ± 0.6‰ to −23.0 ± 0.4‰ over the pH range from 2 to 12. A slight increase in ε values was observed with increasing pH, indicating that while the effect of pH on carbon isotope fractionation is minor, it is distinct. The trend in ε values (12 > 10 > 8 > 5 > 2) is opposite to the trend in E₀, suggesting that E₀ is likely a key factor in governing both the degradation kinetics and isotope fractionation of TCE by KMnO₄. Reported pathways of TCE degradation under varying pH suggest that cyclic esters are transformed into acyclic esters and subsequently hydrolyzed into carboxylic acid intermediates such as formic, oxalic and glyoxylic acids, intermediates known to be pH-dependent [41]. Thus, both the E₀ and the nature of intermediates may influence ε values across the studied pH range.

Figure 4c demonstrates a linear relationship between pH and ε values. This is a useful basis for selecting appropriate ε values in practical applications of KMnO₄-based TCE remediation under varying pH conditions.

3.3. Anions

Chloride (Cl⁻), bicarbonate (HCO₃⁻), nitrate (NO₃⁻), and sulfate (SO₄²⁻) are ubiquitous inorganic anions in natural waters and can potentially influence KMnO₄-mediated TCE degradation [57,58]. As shown in Figure 5a, these anions, at a concentration of 0.2 M, did not significantly affect the TCE degradation rate. Similar have been reported for KMnO₄ oxidation of other organic contaminants such as hexachlorophene and benzophenone-3 [58].

The corresponding ε values in the presence of these anions were −25.6 ± 0.6‰ (Cl⁻), −25.6 ± 0.6‰ (HCO₃⁻), −23.8 ± 0.4‰ (NO₃⁻), and −24.2 ± 0.5‰ (SO₄²⁻), as shown in Figure 5b. Cl⁻ and HCO₃⁻ yielded ε values similar to the blank (−25.4 ± 0.9‰), indicating minimal impact. However, slightly elevated values in NO₃⁻ and SO₄²⁻ systems that, under acidic conditions (final pH = 3.4), NO₃⁻ and SO₄²⁻ may act as competing oxidants, potentially influencing the oxidation pathway of TCE and hence its isotope signature. The stronger oxidizing ability of NO₃⁻ compared to SO₄²⁻ at the same condition may explain the more pronounced deviation in ε values.

Given that Cl⁻ is both a ubiquitous ion and a degradation product of TCE, its influence was further evaluated across a concentration range. As shown in Figure 5d, increasing Cl⁻ concentration up to 2 M had negligible impact on TCE degradation efficiency, although a slight inhibitory effect was observed. This trend was consistent with previous findings by Poulson and Naraoka [40], who reported a ~30% oxidation rate reduction at 0 to 5 M Cl⁻. This is likely due to a salting-out effect, which reduces TCE solubility and thereby slows its oxidation. The carbon isotope fractionation shown in Figure 5e supports this, where higher Cl⁻ concentration (2 M) led to a change in the carbon isotope effect. Corresponding ε values (Figure 5f) decreased from −25.4 ± 0.9‰ (0 M Cl⁻) to −23.9 ± 0.4‰ (2 M Cl⁻), matching trends previously reported by Poulson and Naraoka [40].

Therefore, in practical application, caution should be taken when selecting ε value for quantitative TCE degradation using isotope analysis, as background ions composition, including naturally occurring salinity or introduced by pH adjustments, can influence both degradation kinetics and isotope fractionation.

3.4. HA Concentrations

The influence of HA on contaminant transformation has been widely studied in both natural and engineered systems [52,59]. HA, a ubiquitous macromolecular organic matter in aquatic environments, can affect reaction rate and degradation efficiency in oxidative processes [59]. As shown in Figure 6a, low HA concentration (0.5 mM) did not impact TCE degradation. However, the higher HA concentrations (5–250 mM), TCE removal efficiency decreased markedly. This likely results from competitive consumption of KMnO₄ by HA, with HA being preferentially oxidized at elevated concentrations, thereby reducing KMnO₄ availability for TCE degradation.

The effect of HA on carbon isotope fractionation was evaluated using the relationship between ε values and HA concentrations. As shown in Figure 6b, from −23.5 ± 0.6‰ at 0.5 mM (similar to the blank, −24.4 ± 0.9‰) to −26.6 ± 1.1‰ at 50 mM HA. This trend reflects the growing competition between HA and TCE for KMnO₄ as HA concentration increases. At low HA concentrations (0.5 mM), excess KMnO₄ can oxidize both HA and all of the TCE [51,60]. However, at higher HA concentrations (5–50 mM), KMnO₄ is partially consumed by HA, altering the oxidation conditions and possibly the reaction pathway for TCE degradation. Figure 6c shows a clear relationship between the ε values and HA concentrations, indicating that representative ε values can be selected for use in situ KMnO₄-based remediation. It may be caused by three aspects: 1. HA contains some oxidizing functional groups, which could degrade TCE [52]. 2. HA acts in a competitive role with TCE during KMnO₄ oxidizing [43], and TCE may react with the reaction products of HA, and the corresponding values may differ from those obtained by direct oxidation of TCE with KMnO₄, which may cause the ɛ values changing. 3. The molar ratio of KMnO₄/TCE decreased with the reaction proceeding, which could result in the ɛ values declining, as we discussed in Section 3.1 in this study. Applying the Rayleigh fractionation equation with the appropriate ε values allows for accurate evaluation of TCE degradation in environments with varying HA concentrations.

4. Conclusions

This study represents the first comparative investigation of carbon isotope fractionation during KMnO₄ degradation of TCE under varying concentrations of KMnO₄, TCE and HA across a pH range of 2 to 12 and in the presence of 0.2 M inorganic ions (Cl⁻, HCO₃⁻, NO₃⁻, and SO₄²⁻) in phosphate buffer.

• The molar ratio n(KMnO₄)/n(TCE) has a subtle yet definitive effect on ε values, which should not be overlooked during isotopic interpretation. We observed that ε values are higher when the molar ratio n(KMnO₄)/n(TCE) exceeds 2, compared to ratios below 2. This effect may be attributed to insufficient KMnO₄ (n(KMnO₄)/n(TCE) < 2) being entirely consumed by the excess TCE, generating a substantial amount of MnO₂ colloids that further react with residual TCE. Moreover, under severely limited conditions (n(KMnO₄)/n(TCE) < 1), KMnO₄ primarily cleaves the carbon-carbon double bond of TCE, resulting in the formation of carboxylic acid intermediates. These findings differ from earlier studies, which reported no influence of the n(KMnO₄)/n(TCE) ratio on isotope fractionation.
• A linear relationship was observed between pH and the corresponding ε values, exhibiting a trend opposite to E₀ as pH increased. This suggested that E₀ and TCE degradation intermediates are likely key factors influencing the observed isotope effects during KMnO₄ oxidation.
• The addition of SO₄²⁻ and NO₃⁻ slightly influenced TCE carbon isotope fractionation, likely due to pH reductions during the reaction, which impart mild oxidizing characteristics to these anions under acidic conditions. Increasing initial Cl⁻ concentration from 0 to 2 M also led to elevated ε values, potentially due to the salting-out effect that reduces TCE solubility in the aqueous phase.
• High concentration of HA (50–250 mM) significantly influenced carbon isotope fractionation during TCE degradation, likely due to competition between HA functional groups and TCE for KMnO₄ consumption.

Furthermore, evaluating ISCO performance at field sites may involve considerable uncertainty if the influences of n(KMnO₄)/n(TCE), inorganic ions, HA, and pH on the ε values are disregarded. Therefore, appropriate ε values must be carefully selected for quantitative analysis, and further research is needed to identify the limitations of applying stable carbon isotope analysis at the field scale.