Enhanced Removal of Photosensitive Antibiotics in Water Using CO₂: A Beneficial Exploration of CO₂ Resource Utilization

Abstract

The utilization of carbon dioxide (CO₂) offers an effective approach for alleviating the carbon-reduction pressures associated with fossil energy consumption. However, studies on the use of CO₂ as an auxiliary agent in water treatment to enhance the removal of emerging contaminants are limited. In this study, the photodegradation of ciprofloxacin (CIP) was investigated using ultraviolet (UV) irradiation combined with CO₂ dosing (UV/CO₂). The results demonstrated that the UV/CO₂ system effectively degraded CIP, with CO₂ concentration and solution pH exerting a critical influence. Inorganic anions and metal cations had negligible effects on CIP degradation efficiency, whereas natural organic matter (NOM) had a pronounced inhibitory effect. Mechanistic analysis revealed that superoxide radicals ($·{\mathrm{O}}_2^{-}$) and carbonate radicals (${\mathrm{CO}}_3^{•-}$) were the primary oxidizing species, whereas the excited triplet state of CIP (³CIP*) and singlet oxygen played crucial roles in initiating radical generation. LC–MS analysis and density functional theory calculations indicated that the main degradation routes involved defluorination, decarboxylation, and epoxidation of the piperazine ring. Toxicity assessment indicated that the transformation products generated by UV/CO₂ were less toxic than the parent compound. Furthermore, the UV/CO₂ process demonstrated high energy efficiency, with a low electrical energy per order (EEO) value of 0.4193 kWh·m⁻³·order⁻¹. These findings suggest that the UV/CO₂ system is a promising alternative for the treatment of photosensitive organic pollutants and provides a beneficial pathway for CO₂ utilization.

1. Introduction

The increasing consumption of fossil fuels has led to severe carbon emissions, intensifying global climate change and environmental degradation. As an abundant greenhouse gas and potential carbon resource, carbon dioxide (CO₂) has attracted growing attention for its rational utilization. In recent years, the concept of CO₂ resource utilization has expanded beyond traditional approaches, such as carbon capture, storage, and conversion into fuels or chemicals, to include its application as a functional reagent in environmental remediation. Such strategies not only contribute to carbon mitigation but also create new opportunities for pollution control.

Antibiotics have been extensively used in aquaculture to prevent bacterial illnesses and stimulate growth. However, a large proportion of these compounds is excreted unmetabolized, and residues from feed and fish feces are ultimately discharged into the surrounding waters [1]. Conventional treatment methods, such as biological processes, are not tailored for eliminating antibiotics, resulting in their persistence in aquatic environments [2,3]. The stability and ecological risks of antibiotics highlight the urgent need for efficient and sustainable treatment strategies.

Advanced oxidation processes (AOPs) have attracted considerable attention for antibiotic abatement, as they rely on the in situ generation of reactive oxygen species (ROS), particularly hydroxyl radicals (•OH), which can rapidly and indiscriminately mineralize antibiotic contaminants [4,5,6,7]. Among them, ultraviolet (UV)-based AOPs have attracted particular attention because of their ability to degrade antibiotics while simultaneously reducing antibacterial activity and toxicity [8,9,10]. However, widely studied UV-based systems, such as UV/H₂O₂, UV/persulfate, and UV/chlorination, still face critical limitations, including high chemical costs, regulatory restrictions, poor selectivity in complex matrices, and the potential formation of toxic byproducts [11,12,13]. These drawbacks underscore the need for novel UV-based oxidation processes that are cost-effective, selective and environmentally benign.

Recent investigations have emphasized the significance of carbonate radicals (${\mathrm{CO}}_3^{•-}$) as selective oxidants [14,15]. Unlike non-selective species such as hydroxyl radicals (•OH) and sulfate radicals (${\mathrm{SO}}_4^{•-}$), ${\mathrm{CO}}_3^{•-}$ preferentially reacts with electron-rich functional groups in organic pollutants, with reported second-order rate constants ranging from 10³ to 10⁹ M⁻¹·s⁻¹ [16,17,18]. It reacts efficiently with compounds containing nitrogen, sulfur, or indole moieties, as well as phenolic structures, primarily via electron transfer or hydrogen abstraction pathways [19,20,21]. In UV/H₂O₂ and UV/persulfate systems, the steady-state concentration of ${\mathrm{CO}}_3^{•-}$ has been observed to exceed that of ·OH and ${\mathrm{SO}}_4^{•-}$ by 2–4 orders of magnitude [22]. Furthermore, in natural aquatic environments, indirect photolysis of dissolved organic matter in the presence of bicarbonate/carbonate ions can also yield ${\mathrm{CO}}_3^{•-}$ [23,24]. Collectively, these insights highlight the promise of ${\mathrm{CO}}_3^{•-}$-based advanced oxidation processes as selective and sustainable approaches for the degradation of contaminants. Previous studies have extensively documented the formation and reactivity of ${\mathrm{CO}}_3^{•-}$ in conventional UV-based systems, such as UV/H₂O₂, UV/persulfate, and photo-Fenton systems. While these findings elucidate both the inhibitory and promotional roles of bicarbonate/carbonate in radical-mediated oxidation, the application of these processes generally requires the addition of external oxidants or carbonate salts, leading to high reagent costs and the potential risk of undesired by-products.

Given that dissolved CO₂ equilibrates in water to form H₂CO₃, ${\mathrm{HCO}}_3^{-}$, and ${\mathrm{CO}}_3^{2-}$ depending on the pH [25,26], CO₂ can serve as a precursor for ${\mathrm{CO}}_3^{•-}$ production. This insight provides a unique opportunity to couple wastewater treatment with CO₂ utilization. Considering the rapid increase in anthropogenic CO₂ emissions—from 16 gigatons in 1970 to 37.5 gigatons in 2018—and the significant contribution of wastewater treatment plants (1–2% of global CO₂ emissions) [27], CO₂ valorization in AOPs could simultaneously mitigate greenhouse gas emissions and enhance pollutant removal. Although studies have reported the use of CO₂ as a carbon source to support photosynthetic microorganisms (algae/cyanobacteria) in enhancing bioreactors [28], its application as an auxiliary agent in photochemical wastewater treatment systems has not been reported.

In this context, we propose a novel UV/CO₂ system for the degradation of ciprofloxacin (CIP), a widely used fluoroquinolone antibiotic and a representative contaminant frequently detected in aquaculture effluents. The objectives of this study are to: (i) evaluate the feasibility of UV/CO₂ for efficient CIP abatement; (ii) investigate the influence of key operating parameters and relevant water matrix constituents; (iii) elucidate the underlying reaction mechanism with particular emphasis on the role of ${\mathrm{CO}}_3^{•-}$; and (iv) identify the transformation pathways and assess the ecological toxicity of the intermediates. By integrating pollutant removal with CO₂ resource utilization, this study not only provides new insights into the treatment of photosensitive antibiotics in aquaculture-related waters but also offers a sustainable strategy that aligns water quality improvement with carbon reduction goals.

2. Materials and Methods

2.1. Chemicals and Materials

All reagents employed in this study were of analytical grade and used directly without additional purification steps. Carbon dioxide (99.5%) and nitrogen (99.99%) were supplied by the Zhejiang Outson Gas Co., Ltd. (Hangzhou, China). Ciprofloxacin (CIP), L-histidine (L-His), and indole were obtained from Aladdin Industrial Co., Ltd. (Shanghai, China). Methanol (MeOH), isopropanol (IPA), 1,4-benzoquinone (p-BQ), sodium hydroxide (NaOH), hydrochloric acid (HCl), and phenol were obtained from Sinopharm Chemical Reagent Co., Ltd. (China). Sodium borate buffer, sodium thiosulfate (Na₂S₂O₃), sodium chloride (NaCl), sodium sulfate (Na₂SO₄), sodium nitrate (NaNO₃), sodium dihydrogen phosphate (NaH₂PO₄), 2,4-hexadienoic acid (HDA), and humic acid (HA) were purchased from Macklin Biochemical Technology Co., Ltd. (Shanghai, China). Phenolphthalein was supplied by TCA Kasei Industry Development Co., Ltd. (Shanghai, China). Spin-trapping agents 5,5-dimethyl-1-pyrroline-N-oxide (DMPO) and 2,2,6,6-tetramethylpiperidine (TEMP) were obtained from Sigma-Aldrich (Shanghai, China).

2.2. CIP Degradation Experiments

The experiments were performed in a customized UV photoreactor equipped with a 10 W UV₂₅₄ lamp (Heraeus, Hanau, Germany), as depicted in Figure S1 in the Supplementary Materials. The average irradiance of the lamp was measured as 2.05 mW·cm⁻² using the atrazine chemical actinometry method [29]. A circulating water bath was connected to the reactor to maintain the solution at a constant temperature. Prior to each run, the UV lamp was preheated for 10 min to stabilize its output power. A saturated CO₂ stock solution was prepared by continuously bubbling high-purity CO₂ gas (99.5%) into deionized water at 4 °C, ensuring accurate control of CO₂ dosage. Different CO₂ dosages were achieved by adding varying volumes of saturated CO₂ solution to the reaction system. The headspace above the reaction solution was filled with nitrogen to prevent interference from the atmospheric CO₂. The reaction pH was regulated using 0.2 M sodium borate buffer. Continuous mixing was maintained via magnetic stirring at 1000 rpm. At predetermined intervals (0, 1, 3, 5, 10, 20, and 30 min), 1 mL aliquots were collected and immediately quenched by transferring them into vials containing 50 μL of 1 M sodium thiosulfate. All experiments were conducted in triplicate, and the data presented in this manuscript reflect the average values.

2.3. Analysis Methods

The dissolved CO₂ concentration was quantified using an alkali titration method with phenolphthalein as the visual indicator [30]. Each measurement was performed twice, and the arithmetic mean was calculated. In aqueous systems, several carbonate species (CO₂, H₂CO₃, ${\mathrm{HCO}}_3^{-}$, ${\mathrm{CO}}_3^{2-}$) coexist; their total concentration is collectively denoted as CO_2(T). Ciprofloxacin (CIP) levels were analyzed using a high-performance liquid chromatography (HPLC) system (Agilent 1200, Santa Clara, CA, USA) fitted with an Eclipse XDB-C18 column (4.6 mm × 250 mm, 5 μm). The transformation products were identified using liquid chromatography–mass spectrometry (LC–MS, UPLC/TQ-XS, Agilent 6420, USA). For free radical characterization, DMPO and TEMP were employed as trapping agents, and the resulting spin adduct signals were monitored using an electron paramagnetic resonance spectrometer (Bruker EMXplus 9.5/12). UV–Vis spectrophotometry was used to record the absorption spectra of CIP and its intermediates, and excitation–emission matrix (EEM) fluorescence spectroscopy was used as a complementary technique to trace the degradation process.

3. Results and Discussion

3.1. Influence of Operating Parameter on CIP Degradation by UV/CO₂ Process

3.1.1. Effect of pH Value

Figure 1 illustrates the impact of pH on CIP degradation in both UV and UV/CO₂ processes. Overall, the two systems exhibited comparable pH-dependent trends. As the solution pH increased from 5 to 10, the apparent first-order rate constant (k_obs) increased from 0.0104 to 0.0865 min⁻¹ in the UV system and from 0.0123 to 0.0814 min⁻¹ in the UV/CO₂ system. Similar pH-related behaviors have been documented for direct VUV irradiation and solar photolysis simulations [31,32,33]. Interestingly, in the neutral range (pH 6–8), the UV/CO₂ process exhibited a more pronounced pH influence on CIP degradation than the UV-only process. This difference can be attributed to CIP speciation under varying pH conditions. Protonated species (CIPH⁺) dominate in acidic media (Figure S2a in the Supplementary Materials), whereas neutral or zwitterionic species prevail near circumneutral pH values (7–8). Under alkaline conditions, the deprotonated anion (CIP⁻) becomes predominant [10]. The absorption spectra of these ionization states also differ (Figure S2b in the Supplementary Materials); specifically, the absorbance of CIP at 254 nm gradually increased with increasing pH. Furthermore, in the UV/CO₂ system, ${\mathrm{CO}}_3^{•-}$ can be generated, and the deprotonated CIP species, favored at higher pH, are likely to react with these radicals with enhanced rate constants.

3.1.2. Effect of Initial Concentration of CO₂

Figure S3 of the Supplementary Materials shows the distribution of carbonate species across different pH values. At an acidic pH (<5), H₂CO₃ is the dominant form, whereas ${\mathrm{HCO}}_3^{-}$ prevails in the circumneutral range (pH 6–10), and carbonate ions (${\mathrm{CO}}_3^{2-}$) become predominant under alkaline conditions (pH > 10) [34]. Previous investigations have demonstrated that both ${\mathrm{HCO}}_3^{-}$ and ${\mathrm{CO}}_3^{2-}$ can suppress the photodegradation of organic contaminants because of their strong capacity to scavenge reactive radicals [35,36]. In contrast, Figure 2 highlights that increasing CO₂ dosage exerts a positive effect on CIP degradation in the UV/CO₂ process. At a low dosage (0.5 mM), the effect was negligible. However, when the concentration was increased from 0.5 to 3 mM, k_obs increased from 0.0410 to 0.0945 min⁻¹. Further elevation to 5 mM did not yield any additional improvements, indicating a plateau. These findings suggest that CO₂ supplementation promotes CIP removal, most likely by enhancing the steady-state concentration of ${\mathrm{CO}}_3^{•-}$, which exhibit preferential reactivity toward the aziridine and aromatic moieties of CIP [19,21], thereby accelerating its degradation.

3.1.3. Effect of Initial Concentration of CIP

Figure 3 illustrates the effect of the initial concentration of CIP on its degradation kinetics. When the concentration was increased from 5 to 50 μM, the apparent rate constant decreased significantly from 0.1274 to 0.0509 min⁻¹. Comparable concentration-dependent behaviors have also been observed in other AOPs for the removal of organic contaminants [31,37]. The reduction in degradation efficiency can be ascribed to the intensified competition between CIP molecules and reactive oxidizing species as the pollutant levels increased. Moreover, higher CIP concentrations promoted the generation of intermediate products, which further consumed reactive species and thus competed with the parent compound.

3.1.4. Effect of Temperature

Figure 4 shows the influence of the reaction temperature on CIP degradation. When the temperature was raised from 5 to 30 °C, k_obs increased from 0.0822 to 0.1007 min⁻¹. In contrast, a further increase to 50 °C caused k_obs to decrease to 0.0835 min⁻¹. According to the Arrhenius relationship, elevated temperatures typically accelerate the reaction kinetics. Nevertheless, temperature variation also affects the solubility of dissolved gases such as O₂ and CO₂ in water [38], which are crucial for producing reactive species (e.g., singlet oxygen (¹O₂) and ${\mathrm{CO}}_3^{•-}$) in photosensitized degradation pathways. Moreover, the UV source employed in this work was a low-pressure-mercury lamp, which operates optimally within a limited temperature window [39]; deviations from this range reduce lamp efficiency. Therefore, the non-monotonic-temperature dependence observed here can be explained by the combined effects of intrinsic reaction kinetics, gas solubility changes, and lamp performance.

3.2. Influence of Water Matrices

Common background constituents in natural waters, such as chloride (Cl⁻), nitrate (${\mathrm{NO}}_3^{-}$), and natural organic matter (NOM), are known to influence pollutant degradation in UV-driven-advanced oxidation processes by competing for both photons and reactive oxidants [40]. Hence, evaluating the impact of these matrix components on CIP removal in the UV/CO₂ system is of particular importance. Within the concentration range of 0.5–3 mM, chloride, sulfate, nitrate, bromide, and calcium ions showed negligible effects on CIP degradation efficiency, whereas NOM exhibited a pronounced inhibitory effect (Figure 5). This inhibition can be attributed to the strong UV absorbance of humic acid (HA) in the 200–400 nm range [21], which allows it to act as an effective light-screening agent. In addition, HA has been reported to scavenge ROS [21], further suppressing CIP degradation.

3.3. Degradation Mechanism

3.3.1. Identification of Reactive Species Contributing to CIP Degradation

To clarify the contribution of ROS to CIP degradation within the UV/CO₂ process, selective scavengers were introduced in excess to quench the targeted species. IPA was applied to scavenge •OH (k(IPA +·OH) = 1.9 × 10⁹ M⁻¹·s⁻¹, [41]); P-BQ was employed for scavenging $·{\mathrm{O}}_2^{-}$ (k(P-BQ +$·{\mathrm{O}}_2^{-}$) = 4 × 10⁷ to 10⁸ M⁻¹·s⁻¹, [42]); and L-His was utilized to quench ¹O₂ (k(L-his + ¹O₂) = (2.4–3) × 10⁹ M⁻¹·s⁻¹, [43]). PhOH served as a dual scavenger targeting both •OH and ${\mathrm{CO}}_3^{•-}$, with rate constants of 6.0 × 10⁸ and 2.2 × 10⁷ M⁻¹·s⁻¹, respectively [21]. Additionally, indole was chosen as a more specific scavenger for ${\mathrm{CO}}_3^{•-}$ (k(Indole + ${\mathrm{CO}}_3^{•-}$) = 3.5 × 10⁸ M⁻¹·s⁻¹, [19]). As shown in Figure 6a, the CIP degradation rate decreased after the addition of the various quenchers. The addition of IPA resulted in only a 1.59% reduction, suggesting that •OH played a minor role in this system. In contrast, the addition of L-his led to a 9.93% decrease, indicating the involvement of ¹O₂. More pronounced inhibition was observed with P-BQ, indole, and phenol, which reduced the degradation rate by 52.66%, 68.21%, and 70.15%, respectively, revealing the significant contributions of $·{\mathrm{O}}_2^{-}$ and ${\mathrm{CO}}_3^{•-}$. Furthermore, the difference between the removal efficiencies of indole and phenol (1.94%) was comparable to the decrease observed with IPA, further confirming the minor role of •OH. In summary, within the UV/CO₂ system, ${\mathrm{CO}}_3^{•-}$ likely acted as the dominant reactive species, while $·{\mathrm{O}}_2^{-}$ and ¹O₂ played secondary roles, and •OH made only a limited contribution. It is also important to note that the reactor headspace was purged with N₂ to prevent the intrusion of atmospheric CO₂, and the mineralization of CIP and its intermediates was extremely limited (<2%, Figure S4 in the Supplementary Materials). These observations support the conclusion that the added CO₂ is the principal precursor of ${\mathrm{CO}}_3^{•-}$.

Although ${\mathrm{CO}}_3^{•-}$ plays a crucial role in CIP degradation, it is scarcely generated via direct UV/${\mathrm{CO}}_3^{2-}$ or UV/${\mathrm{HCO}}_3^{-}$ processes [44]. It is well established that UV irradiation of NOM produces excited-state NOM capable of transferring electrons to generate ROS, such as ¹O₂, $·{\mathrm{O}}_2^{-}$, and •OH [45,46]. By analogy, CIP can also be photoexcited to its singlet excited state ¹CIP* (Equation (1)) [47,48]. The ¹CIP* may subsequently undergo three possible pathways: (i) radiative fluorescence, (ii) intersystem crossing (ISC) to the triplet state ³CIP* (Equation (2)), and (iii) direct transformation into photoproducts (Equations (3) and (4)). HDA is a common probe and quencher for triplet-excited chromophores [49]. As shown in Figure 6b, the addition of HDA (10 µM–1 mM) markedly inhibited CIP degradation by 63.3–82.1% in the UV/CO₂ system and by 42.5–56.9% in the UV system, indicating that ³CIP* is a key intermediate. In the presence of O₂, ³CIP* can generate $·{\mathrm{O}}_2^{-}$ and ¹O₂, and may also react with ${\mathrm{HCO}}_3^{-}$/${\mathrm{CO}}_3^{2-}$ to form ${\mathrm{CO}}_3^{•-}$ (Equations (5)–(8)) [31,32,50,51,52]. These ROS ($·{\mathrm{O}}_2^{-}$, ¹O₂, ${\mathrm{CO}}_3^{•-}$) can directly oxidize CIP or further yield •OH [48,53], thereby enhancing degradation. However, ³CIP* can also undergo physical quenching with O₂, forming photoproducts without contributing to the degradation (Equation (9) and Figure S5 in the Supplementary Materials). The conceptual mechanism for CIP photolysis in the UV/CO₂ system is illustrated in Figure S6 of the Supplementary Materials.

$$\mathrm{CIP}+\mathrm{hv} \to {}_{}{}^1\mathrm{C}{\mathrm{IP}}^{*}$$

(1)

$${}_{}{}^1\mathrm{C}{\mathrm{IP}}^{*} \stackrel{\mathrm{ISC}}{\to } {}_{}{}^3\mathrm{C}{\mathrm{IP}}^{*}$$

(2)

$${}_{}{}^1\mathrm{C}{\mathrm{IP}}^{*} \to \mathrm{products}$$

(3)

$${}_{}{}^3\mathrm{C}{\mathrm{IP}}^{*} \to \mathrm{products}$$

(4)

$${}_{}{}^3\mathrm{C}{\mathrm{IP}}^{*}{+\mathrm{O}}_2 \to \mathrm{CIP}+{}_{}{}^1\mathrm{O}_2$$

(5)

$${}_{}{}^3\mathrm{C}{\mathrm{IP}}^{*}{+\mathrm{O}}_2 \to { \mathrm{CIP}}^{·+}{+\mathrm{O}}_2^{·-}$$

(6)

$${}_{}{}^3\mathrm{C}{\mathrm{IP}}^{*}{+\mathrm{CO}}_3^{2-} \to { \mathrm{CO}}_3^{·-}{+\mathrm{CIP}}^{·-}$$

(7)

$${}_{}{}^3\mathrm{C}{\mathrm{IP}}^{*}{+\mathrm{HCO}}_3^{-} \to { \mathrm{H}}^{+}{+\mathrm{CO}}_3^{·-}{+\mathrm{CIP}}^{·-}$$

(8)

$${}_{}{}^3\mathrm{C}{\mathrm{IP}}^{*} {+\mathrm{O}}_2 \to { \mathrm{CIP}+\mathrm{O}}_2$$

(9)

As summarized above, ciprofloxacin (CIP) degradation in the UV/CO₂ process can occur through several reactive pathways, including •OH, ¹O₂, $·{\mathrm{O}}_2^{-}$, ${\mathrm{CO}}_3^{•-}$, and direct photolysis by UV irradiation. To verify the participation of these short-lived-intermediates, electron spin resonance (ESR) spin-trapping-experiments were conducted (Figure 6c,d). When DMPO was applied as the spin trap, signals corresponding to $·{\mathrm{O}}_2^{-}$, ${\mathrm{CO}}_3^{•-}$, and •OH could be detected through the formation of DMPO–$·{\mathrm{O}}_2^{-}$, DMPO–${\mathrm{CO}}_3^{•-}$, and DMPO–•OH adducts, respectively [54,55,56]. For ¹O₂, TEMP was used, producing a characteristic TEMPO triplet spectrum [57]. As shown in Figure 6d, no DMPO–•OH quartet (1:2:2:1) was observed, consistent with the scavenger results, indicating the negligible contribution of •OH. After the addition of methanol, overlapping ESR signals attributed to DMPO–$·{\mathrm{O}}_2^{-}$ and DMPO–${\mathrm{CO}}_3^{•-}$ were recorded. Because the hyperfine constants of the DMPO–${\mathrm{CO}}_3^{•-}$ adduct (αN = 14.35 G, αH_β = 10.7 G, αH_γ = 1.38 G) are nearly identical to those of DMPO–$·{\mathrm{O}}_2^{-}$ (αN = 14.3 G, αH_β = 11.2 G, αH_γ = 1.3 G), the spectra overlapped, complicating the distinction between the two species [34,58]. Upon the simultaneous addition of methanol and P-BQ, the DMPO–${\mathrm{CO}}_3^{•-}$ signal became more discernible, confirming the coexistence of ${\mathrm{CO}}_3^{•-}$ and $·{\mathrm{O}}_2^{-}$. The ESR spectrum obtained using TEMP as the trapping agent (Figure 6c) clearly exhibited a 1:1:1 triplet pattern, further evidencing the presence of ¹O₂. In conclusion, ESR analyses were consistent with radical quenching experiments, collectively demonstrating that ${\mathrm{CO}}_3^{•-}$, $·{\mathrm{O}}_2^{-}$, and ¹O₂ are the major ROS in CIP degradation under UV/CO₂, whereas •OH plays an insignificant role in this process.

3.3.2. Transformation Products

LC–MS was used to characterize the transformation intermediates formed during CIP degradation (Figure S7 in the Supplementary Materials), and the corresponding molecular structures and mass spectra are provided in Table S1 of the Supplementary Materials and Figure 7a. To gain deeper insight into the degradation mechanism and pinpoint the reactive positions within the CIP molecule, density functional theory (DFT) calculations were performed. As shown in Figure 7b, oxygen-containing groups and fluorine atoms were predicted to be the primary nucleophilic sites, whereas the amino group was the main electrophilic center. The HOMO orbital of CIP exhibited electron delocalization over the nitrogen-containing heterocycle and fluorinated ring, indicating a strong electron-donating ability at these sites and their susceptibility to nucleophilic attack (Figure 7c). In contrast, the LUMO orbital was mainly distributed on the ketone-containing heterocycle and fluorinated ring, suggesting that these regions preferentially accept electrons and are vulnerable to electrophilic attacks (Figure 7d). As a nucleophilic species, $·{\mathrm{O}}_2^{-}$ can target both ketone-containing and fluorinated rings [59]. However, the specific atomic sites prone to attack cannot be precisely determined by orbital analysis alone. Fukui function evaluation was used as a theoretical tool to identify molecular sites susceptible to electrophilic (${\mathrm{f}}^{-}$), nucleophilic (${\mathrm{f}}^{+}$), and radical (${\mathrm{f}}^0$) attacks [60,61]. The relevant results are presented in Table S2 of the Supplementary Materials.

Prior investigations suggest that the possible transformation routes of CIP include oxidation at both the piperazine and pyridine moieties, hydroxylation, and removal of fluorine atoms [9,62]. The integration of the observed intermediates with density functional theory data led to the proposal of four mechanistic pathways, as detailed in Figure 7e. Pathway 1 represents a decarboxylation process in which the carboxyl group of CIP is eliminated, yielding intermediate CIP I-1. Subsequent DFT analyses of CIP I-1 (Figure S8 and Table S3 in the Supplementary Materials) suggested that this compound could further undergo defluorination and oxidation at the piperazine ring via free radical attack, leading to the formation of CIP I-2 and CIP I-3. In addition, Liu et al. [63] used the Laplacian Bond Order (LBO) approach to identify potential bond dissociation sites, showing that the F–C bond was the most labile and, therefore, the most prone to cleavage. Consistently, the C17 and F18 atoms of CIP showed relatively high ${\mathrm{f}}^{-}$ and ${\mathrm{f}}^0$ values (Table S2 in the Supplementary Materials), implying that they are likely targets for oxidation by electrophilic species and free radicals. Pathway 2 was characterized as a defluorination route, during which the substitution of the fluorine atom in CIP with a hydroxyl moiety generated the intermediate CIP II-1. This was followed by the substitution of the hydrogen on the piperazine and pyridine rings with a hydroxyl group, yielding CIP II-2 and CIP II-3, respectively. According to the DFT calculation results for CIP II-3 (Figure S9 and Table S4 in the Supplementary Materials). At the identical position of the piperazine moiety, dehydroxylation followed by oxidative transformation gives rise to intermediates CIP II-4 and CIP III-4. CIP II-4 was further oxidized at the pyridine ring to form CIP II-5. Pathway 3 is characterized by an initial oxygen substitution at the piperazine ring, producing CIP III-1, after which hydroxyl substitution of the fluorine atom yields CIP III-2. Subsequently, the free radicals continuously oxidized the piperazine ring to produce CIP III-3, CIP III-4, and CIP III-5. Fukui function analysis revealed that N19, N22, C21, and C23 possess comparatively higher ${\mathrm{f}}^{-}$ and ${\mathrm{f}}^0$ values, highlighting their vulnerability to electrophilic free radical attack and anodic oxidation processes [64]. Pathway 4 was initiated by epoxidation of the piperazine ring, during which ¹O₂ and $·{\mathrm{O}}_2^{-}$ attacked this moiety in CIP. This process triggered a ring-opening reaction that eliminated two hydrogen atoms and introduced two oxygen atoms, yielding CIP IV-1 [9]. Subsequently, the substituents on the piperazine ring of CIP IV-1 underwent further oxidative transformation, and the compound was converted to CIP IV-2 through the loss of a –C=O group. Based on the DFT analyses of CIP IV-2 and CIP IV-5 (Figures S10 and S11; Tables S5 and S6 in the Supplementary Materials), the branched chain of CIP IV-2 was subsequently attacked by reactive species such as ${\mathrm{CO}}_3^{•-}$ and $·{\mathrm{O}}_2^{-}$, yielding intermediates CIP IV-3 and CIP IV-5. These intermediates underwent further oxidative transformation in the pyridine ring, generating CIP IV-6 and CIP IV-7. In most of the identified transformation products, both piperazine and pyridine functionalities underwent substantial oxidation and bond cleavage. Considering the well-documented high reactivity of ${\mathrm{CO}}_3^{•-}$ toward nitrogen-containing and heteroaromatic structures [19,21], it is reasonable to infer that ${\mathrm{CO}}_3^{•-}$ served as the dominant oxidant in the degradation process.

3.4. Toxicity Assessment of Transformation Products

The ECOSAR model was employed to estimate the acute and chronic toxicities of organic contaminants toward fish, Daphnia, and green algae, following the classification criteria outlined in the Globally Harmonized System (GHS) of Classification and Labelling of Chemicals. According to GHS standards, toxicity levels were categorized into four classes: very toxic (Log(LC₅₀/EC₅₀/ChV) ≤ 0), toxic (1 < Log(LC₅₀/EC₅₀/ChV) ≤ 1), harmful (1 < Log(LC₅₀/EC₅₀/ChV) ≤ 2) and not harmful (Log(LC₅₀/EC₅₀/ChV) > 2) [65,66]. Acute toxicity was assessed using the 96 h LC₅₀ for fish, 48 h LC₅₀ for Daphnia, and 96 h EC₅₀ for green algae (Figure 8a). Several intermediates (e.g., I-1 and IV-7) exhibited lower Log(LC₅₀/EC₅₀) values than CIP, suggesting enhanced acute toxicity; however, most intermediates displayed higher values, implying that acute toxicity generally decreased during the degradation process. Chronic toxicity was evaluated using the ChV values (Figure 8b). While a subset of intermediates showed Log(ChV) < 0, reflecting elevated chronic toxicity and warranting attention, most intermediates demonstrated Log(ChV) > 1, indicating limited long-term ecological risk. Collectively, these findings highlight that the UV/CO₂ system not only facilitates the removal of CIP but also mitigates its potential environmental hazards.

3.5. Actual Water Testing and Economic Analysis

FQs are widely distributed in aquatic systems and have been detected in municipal wastewater effluents [67], lake water [68], and marine environments [69]. To investigate the practical feasibility of the UV/CO₂ process for CIP elimination, three representative water matrices—lake water (LW), secondary sedimentation tank effluent (SSTW), and seawater (SW)—were selected for evaluation. As illustrated in Figure 9, after 30 min of irradiation, the CIP removal efficiencies were 88.9% in LW, 83.7% in SSTW, and 72.8% in SW, all of which were lower than that achieved in ultrapure water. The results from the 3D-EEM spectra (Figure S12 in the Supplementary Materials), together with the water quality parameters (Table S7 in the Supplementary Materials), indicated that SSTW contained the highest concentration of dissolved organic matter, whereas LW and SW exhibited comparatively lower levels. The presence of organic matter was identified as the major inhibitory factor, primarily due to its radical-scavenging capacity, which consequently suppressed CIP degradation.

To assess the energy efficiency of the UV/CO₂process, a kinetic model was combined with the Electrical Energy per order (EE/O) approach. EE/O (kWh·m⁻³·order⁻¹) represents the electrical energy required to achieve a one-order-of-magnitude decrease in pollutant concentration per cubic meter of treated water [70]. The energy demand of the UV irradiation system (EE/O_UV) and the contribution from CO₂ consumption (CO₂/O) were calculated according to Equations (10) and (11). Considering that the reported energy requirement for industrial CO₂ production typically ranges between 80 and 120 kWh·t⁻¹ [71], a representative value of 100 kWh·t⁻¹ was adopted in this study. The CO₂/O was further expressed as equivalent electrical energy consumption in the same unit (kWh·m⁻³·order⁻¹) as EE/O_UV (Equation (12)). Consequently, the overall energy demand of the UV/CO₂ process (EE/O_total) was determined using Equation (13).

$${\mathrm{EE}/\mathrm{O}}_{\mathrm{UV}}={\displaystyle \frac{\mathrm{P}·\mathrm{t}·1000}{\mathrm{V}·\log \left(\frac{{\mathrm{C}}_{\mathrm{i}}}{{\mathrm{C}}_{\mathrm{f}}}\right)}}={\displaystyle \frac{38.4·\mathrm{P}}{\mathrm{V}·\mathrm{k}}}{ (\mathrm{kWhm}}^{-3}{\mathrm{order}}^{-1})$$

(10)

$${\mathrm{CO}}_2/\mathrm{O}={\displaystyle \frac{{\mathrm{C}}_{{\mathrm{CO}}_2}}{\log \left(\frac{{\mathrm{C}}_{\mathrm{i}}}{{\mathrm{C}}_{\mathrm{f}}}\right)}}{ (\mathrm{mgL}}^{-1}{\mathrm{order}}^{-1})$$

(11)

$${\mathrm{EE}/\mathrm{O}}_{{\mathrm{CO}}_2}{=-1 \times 10}^{-4}{ \times \mathrm{CO}}_2{/\mathrm{O} (\mathrm{kWhm}}^{-3}{\mathrm{order}}^{-1})$$

(12)

$${\mathrm{EE}/\mathrm{O}}_{\mathrm{total}}{=\mathrm{EE}/\mathrm{O}}_{\mathrm{UV}}{+\mathrm{EE}/\mathrm{O}}_{{\mathrm{CO}}_2 }{(\mathrm{kWhm}}^{-3}{\mathrm{order}}^{-1})$$

(13)

where P denotes the UV lamp output power (kW), t is the irradiation duration (h), V represents the effective reactor volume (L), k corresponds to the apparent first-order rate constant (k_obs, min⁻¹), C_i and C_f are the initial and final concentrations of the target compound (mg·L⁻¹), respectively, and C_CO2 refers to the dissolved CO₂ concentration in the reaction medium (mg·L⁻¹).

Under the operating conditions of an initial CIP concentration of 20 µM, CO₂ dosage of 3 mM, temperature of 20 °C, and pH 8.0, the specific electrical energy consumptions were calculated as 0.406, 0.0133, and 0.4193 kWh·m⁻³·order⁻¹ for the UV source, CO₂ utilization, and the combined UV/CO₂ system, respectively. A comparison with the reported EE/O values for other UV-based advanced oxidation processes ([72,73] in Table S8 of the Supplementary Materials) revealed that the UV/CO₂ process required substantially lower energy. Specifically, its energy demand was only approximately 1/50 that of UV/H₂O₂ and 1/10 that of UV/O₃. The notably lower energy requirement demonstrates the superior economic feasibility of the UV/CO₂ process, providing strong evidence of its applicability in practical water treatment scenarios.

4. Conclusions

This study systematically elucidated the degradation of ciprofloxacin (CIP) in the UV/CO₂ process and highlighted the pivotal role of carbonate-derived reactive species. Relative to the UV-alone control (pH adjusted using HCl/NaOH), the UV/CO₂ system exhibited markedly enhanced degradation efficiency across a wide pH range, demonstrating that the improvement arose from reactive carbonate radicals rather than simple pH effects. The process was further validated in real water matrices containing multiple background constituents, confirming its robustness in complex conditions. Quenching tests and EPR analyses revealed that UV irradiation induces the formation of the triplet excited state of CIP (³CIP*), which subsequently reacts with dissolved oxygen and carbonate species to generate oxidants including ¹O₂, $·{\mathrm{O}}_2^{-}$ and ${\mathrm{CO}}_3^{•-}$. These reactive species drive a series of transformation pathways involving decarboxylation, defluorination, oxidation of the piperazine moiety, and pyridone ring modification. Toxicity assessment demonstrated a substantial reduction in ecological risk following UV/CO₂ treatment, underscoring its effectiveness and environmental relevance. Overall, the UV/CO₂ process achieved efficient CIP removal while offering advantages in terms of energy efficiency and sustainability, supporting its potential as a practical alternative for water purification. Nevertheless, further research is warranted to integrate kinetic modeling with radical quantification, investigate complex water matrices in greater detail, and evaluate the long-term ecological impacts of transformation products, thereby promoting the broader application of UV/CO₂ processes in water treatment.