Photocatalytic Degradation of Acetaminophen by g-C₃N₄/CQD/Ag Nanocomposites from Aqueous Media

Abstract

Ternary g-C₃N₄/CQD/Ag photocatalysts were synthesized via deposition of carbon quantum dots (CQDs) and silver nanoparticles (Ag) onto graphitic carbon nitride (g-C₃N₄) for efficient acetaminophen degradation. The nanocomposites exhibited enhanced photoresponse and broad-spectrum photocatalytic activity under both UV (254 nm, 250 W) and Xenon (>420 nm, 500 W) irradiation. Characterization by XRD, FTIR, SEM, PL, and EDX elucidated the material’s composition, structure, morphology, and optical properties. Optimized photocatalytic degradation of acetaminophen (50 mg/L) was achieved at pH 7 with 0.6 g/L catalyst loading and 60 min irradiation, yielding degradation efficiencies of 87.5% (UV) and 85.3% (Xenon). Radical quenching experiments and GC-MS analysis identified hydroxyl radicals as the primary reactive species and revealed a gradual decrease in intermediate toxicity during mineralization. This study demonstrates the superior photocatalytic performance of the ternary g-C₃N₄/CQD/Ag nanocomposites compared to binary systems for effective acetaminophen removal.

1. Introduction

A number of stubborn organic substances, like surfactants, pharmaceuticals, flame retardants, plasticizers, fragrances, and other trace chemical compounds, which are often linked to human illnesses, have been identified in aquatic environments [1]. Recently, pharmaceutical and personal care products (PPCPs) have emerged as pollutants of growing concern because of their negative impacts on ecosystems and human health [2]. Although pharmaceutical products are essential for human health and well-being, they eventually become waste [3]. If not properly disposed of, they can become a global environmental issue. Pharmaceutical contaminants are already deemed as pollutants of emerging concern [3]. Acetaminophen (ACT), a widely used pharmaceutical, is an example of such an emerging contaminant. It is commonly found in over-the-counter and prescription medications to treat fever and pain [4]. Approximately 1.4 × 10⁵ t of ACT are consumed worldwide annually. Because of its great solubility (12.78 g/L) in water and the fact that only 5–15% is absorbed by the body, a large portion is excreted and can enter the environment, contaminating various aquatic systems, including surface water, groundwater, and drinking water [5]. For example, a hazard quotient of 1.8 for ACT has been reported, indicating its potential to cause adverse ecological effects. These facts highlight the urgent need for effective removal of ACT from water [6]. Conventional treatment systems, such as activated sludge or adsorption, are insufficient for completely removing pharmaceutical pollutants. This inadequacy has driven the development of new technologies and materials, including carbon nanomaterials [7]. For instance, Ahmed et al. [8] revealed that solely 2% of participants in Pakistan appropriately disposed of their unused and expired medications, while 83% discarded them with regular garbage—a practice consistent with studies conducted in Brazil, Bangladesh, Afghanistan, and Ireland. g-C₃N₄ has been deemed as a sustainable and effective metal-free photocatalyst for degrading organic contaminants because of its cost-effectiveness, high stability, non-toxicity, and proper band gap (2.7 eV) [9]. Nonetheless, pure g-C₃N₄ faces challenges, like inappropriate electrical conductivity, high electron–hole reintegration rates, and hindered visible light absorption [10]. To enhance its photocatalytic performance, various strategies like doping with noble and transition metals, non-metal doping, and surface modifications have been applied [11]. For instance, Shamilov et al. (2023) [12] investigated the photocatalytic performance of the semiconductor g-C₃N₄ in comparison with the heterostructured composites CdZnS/g-C₃N₄ and ZnIn₂S₄/g-C₃N₄ for the degradation of rhodamine B. Their findings indicated that the rate constant for rhodamine B degradation was 2.75 times higher with CdZnS/g-C₃N₄ and 1.9 times higher with ZnIn₂S₄/g-C₃N₄ compared to pure g-C₃N₄. Similarly, Chebanenko et al. (2024) [13] reported that the incorporation of cobalt oxide into the g-C₃N₄ matrix significantly reduced charge carrier recombination and enhanced the generated photocurrent by a factor of 1.2. Doping g-C₃N₄ with noble metals like silver (Ag) can remarkably promote its photocatalytic effectiveness, as these metals can capture electrons and boost the separation of electron–hole pairs [14]. When g-C₃N₄ is combined with metals, a metal-semiconductor heterojunction is formed, which enhances photocatalytic performance [15]. Ag, in particular, is widely studied for its catalytic, biosensor, and biomedical applications. For instance, silver chloride (AgCl) has appealing properties as a broad band gap semiconductor (3.25 eV). By combining AgCl with g-C₃N₄, the separation of charge carriers can be enhanced under visible light, leading to improved photocatalytic performance [16]. Moreover, carbon dots (CDs), such as carbon quantum dots (CQDs), can serve as electron mediators between g-C₃N₄ and AgCl semiconductors. Although CDs do not possess intrinsic photocatalytic effectiveness, their unique optical and electronic characteristics—like robust light absorption, photoexcited electron transfer capability, up-conversion photoluminescence, tunable bandgap, and inhibition of electron–hole reintegration—make them ideal for constructing highly efficient photocatalytic systems [17]. CQDs, in particular, have been widely used to enhance photocatalytic ability. Studies have shown that incorporating CQDs with g-C₃N₄ narrows the bandgap and significantly improves the degradation rate of organic pollutants [18]. There are few publications on using CQDs/g-C₃N₄ heterojunctions as photocatalysts for removing organic pollutants from wastewater [18]. Therefore, this study evaluates the applicability of efficient g-C₃N₄/CQDs/Ag nanocomposites, aiming to achieve significantly improved photocatalytic effectiveness under UV and Xenon lights for decontaminating acetaminophen. The current work highlights the innovative implementation of carbon dots as UV and Xenon lights-driven photocatalysts. These carbon dots serve as effective co-catalysts, enabling the implementation of the full spectrum of sunlight from the previous reports [19]. Liu et al. (2019) [20] carried out a comprehensive analysis on the incorporation of carbon dots (CDs) into semiconductors like g-C₃N₄. Their findings highlighted several significant enhancements in photocatalytic such as allowing for more efficient application of visible light in photocatalytic processes, minimizing the reintegration of electrons and holes (e⁻/h⁺), which contributes to a longer lifetime of charge carriers, and increasing the available active sites for facilitating more efficient chemical transformations.

2. Material and Methods

2.1. Chemicals and Materials

Acetaminophen (C₁4H₂₂N₂O₃, ≥98.0%), silver nitrate (AgNO₃, 99%), melamine (C₃H₆N₆), sodium chloride (NaCl), sulfuric acid, ascorbic acid, and ethanol were provided from Sigma-Aldrich Co., Saint Louis, MO, USA. Sodium hydroxide (NaOH), Nitric acid (HNO₃), tert-butanol (t-BuOH), benzoquinone (BQ), 1,4-benzoquinone, and sodium azide (NaN₃) were obtained from Merck Co., Darmstadt, Germany. All chemical compounds utilized in the current work were of analytical grade and were utilized as received without additional treatment. Deionized distilled water (DI) was utilized for washing materials and preparing solutions. The chemical structure and other attributes of acetaminophen are presented in

Table 1. Chemical Structure and Characteristics of ACT.

Category	Detail	Category	Detail
Structure		IUPAC name	N-(4-hydroxyphenyl) Acetamide
Molecular mass	151.2 g/mol	Trade names	Tylenol/Panadol
Density	1.263 g/cm3	Melting point	169–172 °C
Log Kow	2	Water solubility	1400–2400 mg/L
		pKa	9.38

[21].

2.2. Fabrication of Nanocomposite

2.2.1. Synthesis of g-C₃N₄

g-C₃N₄ was synthesized through the thermal poly-condensation of melamine under a controlled atmosphere, following a previously reported method [22]. Specifically, 20 g of analytical-grade melamine powder was placed in a covered alumina crucible and heated in a muffle furnace at a ramp rate of 10 °C/min to 550 ± 20 °C for 4 h. The resulting powder was then transferred to an open aluminum crucible and maintained at 500 °C for 2 h with a ramp rate of 1.5 °C/min. After cooling to room temperature, the product was collected and ground into a fine powder, yielding the final light-yellow g-C₃N₄ powder.

2.2.2. Synthesis of Ag/g-C₃N₄ Nanocomposites

Ag/g-C₃N₄ was synthesized using the in situ co-precipitation method. First, 1 g of g-C₃N₄ was dispersed in 50 mL of water under ultrasonic treatment. Then, 25 mg of AgNO₃ was gradually added to the solution and stirred magnetically in the dark for 10 min to facilitate sufficient adsorption of Ag⁺ ions onto the g-C₃N₄ surface [23]. At pH 7, g-C₃N₄ carries a negative surface charge, promoting the adsorption of Ag⁺ ions [24]. Subsequently, 10 mL of NaCl solution (0.01 g in 10 mL of water) was added dropwise to the suspension while stirring for 10 h. The resulting yellow precipitate was collected by centrifugation at 120 rpm, washed twice with water and ethanol, and dried in an oven at 50 °C for 12 h.

2.2.3. Synthesis of CQDs

CQDs were synthesized using a modified hydrothermal method based on previous studies [25]. In a typical procedure, 0.5 g of ascorbic acid and 25 g of ethanol were first dissolved in 25 mL of deionized (DI) water. The resulting mixture was then transferred to a Teflon-sealed autoclave and subjected to a hydrothermal process at 180 °C for 4 h. After naturally cooling to room temperature, the dark brown suspension was centrifuged at 120 rpm for 20 min to remove bulk particles. The obtained CQDs solution was stored at 4 °C for subsequent synthetic procedures

2.2.4. Synthesis of g-C₃N₄/CQD

The g-C₃N₄/CQDs composites were prepared following our previous report [26]. In this process, 1 g of g-C₃N₄ and 50 g of CQDs (with CQDs comprising 2 wt% of the g-C₃N₄ mass) were mixed and stirred for 10 min. The mixture was then transferred into an autoclave and subjected to a reaction at 180 °C for 4 h. After cooling, filtration, and drying, the binary composite photocatalyst g-C₃N₄/CQDs was obtained.

2.2.5. Synthesis of g-C₃N₄/CQDs/Ag

A total of 1 g of the g-C₃N₄/CQD composite was placed in a 200 mL beaker, followed by the addition of 150 mL of deionized water. Subsequently, 0.5 g of g-C₃N₄/Ag was introduced into the beaker. The mixture was heated at 130 ± 10 °C for 4 h. The resulting solid was then transferred to an oven and dried at 70 °C to obtain the g-C₃N₄/CQDs/Ag composite.

2.3. Characterization

The crystal structures of both g-C₃N₄/CQD and g-C₃N₄/CQD/Ag were examined utilizing advanced (XRD) over a 2θ range from 10° to 80° with graphite monochromatic CuKα radiation (λ = 1.541 Å). The chemical functional groups and structures of the as-constricted composites were recorded using (FTIR) spectroscopy., (Model: Spectrum two, PerkinElmer Co, Waltham, MA, United states), covering the range from 400 cm⁻¹ to 4000 cm⁻¹. FESEM and EDX analyses (Sigma VP model, ZEISS Co., Oberkochen, Germany) were carried out to study the prepared photocatalysts. Photoluminescence (PL) spectra (model CARY ECLIPSE. Varian Co, Santa Clara, CA, USA), ranging from 200 nm to 800 nm, were obtained using a fluorescence spectrometer to evaluate the recombination rate of multiple specimens (Instrumental Analysis Lab, Mahamax Industrial and Laboratory Analysis, Tehran, Iran). The content of acetaminophen was quantified by spectrophotometry at a wavelength of 235 nm. The value of total organic carbon (TOC) of specimens was quantified using a Multi N/C 3100 analyzer (Analytik Co., Jena, Germany). The intermediates formed during ACT degradation were tracked adopting gas chromatography-mass spectrometry, model of GC-MS, Agilent 5975C (Agilent Technologies Co., Santa Clara, CA, USA) with an electrospray ionization source in positive ionization mode.

2.4. Photocatalytic Effectiveness Experiments

Photocatalytic oxidation tests were performed in a glass photochemical reactor (15 × 15 × 30 cm, Figure 1) featuring a 10 L continuous-flow water cooling system. For electron excitation, UV (P = 250 W, λ = 254 nm, Philips, Amesterdam, The Netherlands, 123.7 mW/cm², Length = 10 cm) and Xenon (100 mW/cm² using a 500 W Xenon lamp (λ > 420 nm, Philips, Amesterdam, The Netherlands, length = 6 cm) lamps were used separately. In order to homogenize the mixture containing catalyst and pollutant, a mechanical stirrer was used with 40 RPM.To prevent direct contact between the lamps and the reaction solution, and to control the spectral irradiance, a quartz tube (25 cm length, 4 cm diameter) was utilized. A standard optical filter with a 420 nm cut-off wavelength was applied to the quartz tube to selectively filter the Xenon lamp’s output Xenon. The distance between the lamps and the reactor walls was considered to be about 7 cm. Also, photometer device (Hagner EC1-Xwas., Solna, Sweden) used to measure the radiation intensity. A stock solution of acetaminophen (1000 mg/L) was provided via dissolving it in distilled water under magnetic stirring. In the next step, 1000 mL of the catalyst and ACT mixture was added to the reactor according to the design conditions of the experiment. Prior to illumination, the suspension was stirred in the dark for 30 min to achieve adsorption–desorption equilibrium between the contaminants and the catalysts.

After the completion of the photocatalytic process, the experiment was conducted to determine the effect of UV radiation alone in the photolysis of acetaminophen without the effect of nanoparticles. Specimens were collected at predetermined interval times, and the catalyst was removed from the solution using a Whatman 0.45 µm filter membrane. Acetaminophen content in the supernatant was determined utilizing a UV/Vis spectrophotometer (Hach-Lange-DR 5000, Loveland, CO, USA) at λ_max = 235 nm. The tests were performed under various conditions, including different initial pH levels (3, 5, 7, 9, and 11), nanocomposite doses (0.1, 0.2, 0.3, 0.4, 0.5, 0.6, 0.7, and 1 g/L), initial acetaminophen contents (10, 30, 50, and 100 mg/L), and irradiation times (15, 30, 45, 60, 75, 90, 105, and 120 min) [27]. The initial pH level of each media was regulated using 0.1 N HCl or 0.1 N NaOH and quantified with a pH meter (Metron, Herisau, Switzerland). All experiments were conducted under ambient circumstances for 2 h at 25 ± 1 °C with constant stirring.

2.5. Scavenger-Quenching Experiments

Scavenger-quenching experiments are crucial for elucidating the roles of multiple reactive agents in photocatalytic processes. This study utilized various scavengers to assess their impact on the degradation of ACT, including t-Butyl Alcohol (t-BuOH) (14.8 mM), 1,4-Benzoquinone (BQ) (8.8 mM), and Ammonium oxalate (AO) (8.8 mM), which were deployed as hydroxyl radical (°OH), hole (h⁺), and superoxide radicals (°O₂⁻) scavengers, respectively [28]. Subsequently, each suspension sample was quantified by its absorbance at a 235 nm wavelength utilizing a UV-vis spectrophotometer. Quenching in photocatalytic decontamination reactions will be partly suppressed, and the extent of decline reveals the importance of the corresponding degradation agents.

2.6. Statistical Analysis

Data on elimination rate were analyzed using Origin (ver. 2018) software via a one-way analysis of variance experiment. The removal efficiency for each factor was computed using the following equation:

$$\mathrm{r}\mathrm{e}\mathrm{m}\mathrm{o}\mathrm{v}\mathrm{a}\mathrm{l} \mathrm{e}\mathrm{f}\mathrm{f}\mathrm{i}\mathrm{c}\mathrm{i}\mathrm{e}\mathrm{n}\mathrm{c}\mathrm{y}\mathrm{\%}={\displaystyle \frac{\left({\mathrm{C}}_0-{\mathrm{C}}_{\mathrm{e}}\right)\times 100}{{\mathrm{C}}_0}}$$

(1)

where C₀ is the initial content of acetaminophen (mg/L), and C_e is the parameter acetaminophen content in the treated solution after a given time (mg/L).

3. Results and Discussion

3.1. Structure and Composition Photocatalysts

XRD patterns were utilized to determine the structure and phase of the nanomaterials. The XRD patterns of the as-constructed g-C₃N₄ and g-C₃N₄/CQD/Ag materials are displayed in Figure 2a. The peaks at 12.8° and 27.6° for g-C₃N₄ align well with previously reported g-C₃N₄ nanomaterials [20,29]. The most prominent peak at 27.6°, corresponding to the (0 0 2) plane of g-C₃N₄, arises from the stacking of layers in the conjugated aromatic system.

The weaker peak at 12.8° is associated with the layered packing structure of tri-s-triazine units, corresponding to the (1 0 0) plane. For the g-C₃N₄/CQD nanocomposite, a similar pattern was depicted, indicating that the structure of g-C₃N₄ remains unchanged upon introducing CQDs, and no diffraction peaks of CQDs were depicted because of their rare quantity in the material. In the XRD patterns of the g- g-C₃N₄/CQD/Ag nanocomposite, besides the characteristic peaks of g-C₃N₄, peaks corresponding to Ag were detected at 2θ = 32.95° (1 1 1), 38.26° (2 0 0), 57.47° (2 2 0), and 72.5° (3 1 1) planes. The presence of the (0 0 2) plane at 27.6° verifies its retention in the graphitic nature. The formation of the ternary nanocomposites was successful, with no diffraction peaks corresponding to impurities, indicating high purity of the nanocomposites. FTIR technique was undertaken to examine the functional groups present in the as-constructed materials, as displayed in Figure 2b. The pristine g-C₃N₄ material exhibited bands at 1236–1641 cm⁻¹ (C–N heterocycles), 888 cm⁻¹(N-H deformation), and 807 cm⁻¹ (tri-s-triazine ring mode of condensed C–N heterocycles). Additionally, the sharp absorption peak at 810 cm⁻¹ is owing to the bending vibration of triazine units. The spectra of the g-C₃N₄/CQD/Ag composites revealed that all the peaks closely resemble those of pure g-C₃N₄, indicating that the addition of CQDs did not affect the chemical structure of g-C₃N₄. Nonetheless, after Ag doping, the stretching vibration of the C-N heterocycles relocated to a higher frequency region, which could be because of an elevation in the force constant of the C-N heterocycles following the addition of Ag. The microstructure and morphology of the as-obtained specimens were diagnosed via Field Emission Scanning Electron Microscopy. The FE-SEM image of pristine g-C₃N₄ (Figure 3a) exhibited a smooth surface with multilayered sheet-like structures, which aligns well with previous studies [30]. The morphology of the g-C₃N₄ samples remained largely unchanged following CQD modification, indicating that the integration of CQDs did not significantly alter the structural characteristics of g-C₃N₄. However, after the deposition of silver nanoparticles, the g-C₃N₄/CQDs/Ag photocatalyst displayed a smoother and more uniform surface, with Ag nanoparticles evenly distributed across the g-C₃N₄ sheets, as depicted in Figure 3b. Notably, the size of the deposited silver nanoparticles increased proportionally with the silver content. Based on Figure 3b, the average diameter of the Ag nanoparticles was measured to be approximately 75.43 nm (D₁ = 88.63 nm, D₂ = 62.24 nm). These findings confirm the successful incorporation of both CQDs and silver nanoparticles onto the g-C₃N₄ surface. EDX analysis in Figure 4a revealed that g-C₃N₄ mainly consists of C and N elements. In addition, C, N, and Ag elements were clearly depicted in the g-C₃N₄/CQD/Ag nanocomposite (Figure 4b). Consistent with the EDX spectra, the presence of Ag was detected, with a content of 13 wt% and an atomic ratio of 2.59% Ag, which was absent in the pristine pure g-C₃N₄ specimen, thereby confirming the successful doping of Ag into g-C₃N₄. The quantitative analyses of the EDX outcomes are summarized in

Table 2. Weight and atomic percentage for different elements in g-C₃N₄ and g-C₃N₄/CQD/Ag using EDX Analysis.

Components	g-C3N4
	Weight %	Atomic %
C	40.57	44.32
N	59.43	55.68
Ag	-	-
Total	100	100
Components	g-C3N4/CQD/Ag
	Weight %	Atomic %
C	35	42.1
N	52	55.3
Ag	13	2.59
Total	100	100

. PL properties were also studied to assess the carrier separation behavior’s contribution to photocatalytic degradation efficiency. As displayed in Figure 4c, pristine g-C₃N₄ exhibited a strong PL emission peak at 600–610 nm. After forming the g-C₃N₄/CQD/Ag composite with a suitable quantity of Ag, the PL emission peak shifted to 570–590 nm, indicating a superior separation of electron–hole pairs and a declined reintegration combination probability of photo-induced electron–hole pairs, which enhances visible light performance.

3.2. Photocatalytic Decomposition of ACT Under Ultraviolet and XenonXenon-Light Illumination

3.2.1. Impact of Media pH

This research examined the impact of initial pH on decontaminating ACT using the g-C₃N₄/CQD/Ag catalyst. Figure 5 illustrates the investigation of the effects of pH variation, photocatalyst composition, pollutant chemical properties, and photocatalyst surface charge on the photocatalytic process. To scrutinize the impact of pH on the surface charge of the g-C₃N₄/CQD/Ag photocatalyst, zeta potential measurements were taken across different pH levels. The point of zero charge (pH_pzc) for the g-C₃N₄/CQD/Ag nanocomposite was determined to be at pH 6.5. The decomposition rate of ACT under Xenon lamp and UV irradiation enhanced as the pH moved from acidic to neutral conditions, but it decreased when the pH was boosted to more alkaline circumstances (pH 11). The pKa of ACT is 9.5, indicating that ACT becomes negatively charged at pH levels above 9.5. When the pH rises (e.g., to 11.0), the electrostatic repulsion between the negatively charged surface of the catalyst and the negatively charged ACT molecules becomes stronger. The observed decrease in the decomposition efficiency at pH levels above 7 can be attributed to the intensified electrostatic repulsion between the negatively charged photocatalyst surface (at pH > pHpzc) and the negatively charged OH⁻ ions, along with the growing abundance of deprotonated, negatively charged ACT molecules [31]. Similarly, Noroozi et al. demonstrated that pH is a significant factor in decontaminating acetaminophen in water environments when using carbon nanocomposites [32]. In their study, the decomposition efficiency elevated from 62% to 100% when the pH was raised from 5 to 7, but diminished as the pH was further increased from 7 to 10. This is because the capability of the catalyst to absorb and decompose contaminants depends on its surface charge state, the type of charge, and its oxidation potential [32]. The results suggest that the observed phenomena can be explained by considering the electrostatic interactions between the ACT molecules and the photocatalyst, along with the ionic state of ACT and the surface charge of the photocatalyst at different pH conditions.

3.2.2. Impact of Oxidation Time

Figure 6 illustrates the decomposition rate of ACT utilizing the g-C₃N₄/CQD/Ag catalyst under Xenon and UV light, achieving 97.7% and 98.2% removal, respectively, within 60 min of the reaction time. After 60 min, no additional enhancement in the elimination rate was observed, probably because of the blockage or overlap of the active sites and surface area of the adsorbents. The incorporation of Ag into the photocatalyst slightly improved the elimination rate of ACT, which may be linked to the enhanced surface area and the number of active sites in the g-C₃N₄/CQD/Ag composite [33]. Cako et al. [34] also investigated the decomposition of ACT using a different photocatalytic system (P,S-g-C₃N₄/2D TiO₂) and found that the decomposition rate of ACT increased with longer contact time. In their study, the degradation of ACT reached 99.5% after 60 min, demonstrating that prolonged contact time enhances the likelihood of interactions between the pollutant and the nanocomposite. Additionally, the active sites on the nanocomposite, which are crucial for pollutant absorption, become more effective over extended periods [34].

3.2.3. Impact of Photocatalyst Content

The influence of the catalyst content on photocatalytic oxidation is displayed in Figure 7. Various contents of the g-C₃N₄/CQD/Ag material, spanning from 0.1 to 1 g/L, were tested with a constant acetaminophen concentration of 50 mg/L. As the photocatalyst content elevated from 0.1 to 0.6 g/L, the decontamination of acetaminophen improved from 64% to 98.3% under xXenon light and from 66% to 99% under UV light. However, further elevating the content from 0.7 to 1 g/L gave rise to a decline in ACT photodegradation. The initial increase in photocatalytic effectiveness may be ascribed by the increasing number of active sites available at higher photocatalyst concentrations, which enhances photocatalytic degradation [35]. However, beyond a certain point, increasing the concentration causes increased solution turbidity, which reduces light penetration and light harvesting. Additionally, the aggregation of catalyst nanoparticles reduces the available active surface area for light absorption, thus decreasing photocatalytic decomposition efficiency [36]. Consequently, higher photocatalyst concentrations reduce the specific activity of the as-constructed photocatalyst. The optimal photocatalyst concentration for maximum ACT photodegradation efficiency with g-C₃N₄/CQD/Ag was found to be 0.6 g/L. Xiong et al. [37] found that increasing the catalyst dosage enhances the specific surface area and the number of active sites available for photocatalytic reactions, which aligns well with the results reported in other studies.

3.2.4. Impact of Initial Acetaminophen Content

The influence of differing initial contents of acetaminophen in the span of 10–100 mg/L on photodecomposition was scrutinized. As shown in Figure 8, for initial ACT concentrations between 10 and 100 mg/L, the photodecomposition approach was rapid, and nearly all ACT was decontaminated after 60 min of Xenon-light and UV illumination. Nonetheless, as the initial ACT content enhanced from 30 to 100 mg/L, a significant diminishing in the degradation rate was observed. At greater contents, the ratio of active sites to contaminant molecules decreases, and the number of radical agents on the photocatalyst surface is reduced. This leads to a relatively smaller number of radicals being accessible for degrading each ACT molecule, thereby diminishing the rate of ACT decomposition [38]. Gotostos et al. [39] found that as the initial content of ACT in the reaction environment rises, more hydroxyl and oxygen radicals are required for the oxidation of ACT. Consequently, under constant test conditions (such as nanocatalyst dose, contact time, and lamp irradiance), the decomposition rate of acetaminophen diminishes with elevating initial content.

3.2.5. Trapping Experiments

The photocatalytic degradation mechanism of the as-fabricated g-C₃N₄/CQD/Ag catalysts was investigated using radical scavenger experiments. Ammonium oxalate (AO), tertiary butyl alcohol (t-BuOH), and benzoquinone (BQ) were employed to capture holes (h⁺), hydroxyl radicals (°OH) and superoxide radicals (°O₂⁻), respectively. Figure 9 illustrates the impact of multiple scavengers on the photocatalytic effectiveness of g-C₃N₄/CQD/Ag catalysts for decontaminating acetaminophen. The findings manifest that the introduction of t-BuOH had the most notable impact on the photocatalytic effectiveness, reducing the degradation rate of ACT from 100% to 49.3% under Xenon irradiation and to 51.2% under UV irradiation. This was followed by BQ, which reduced the degradation rate to 93.2% and 96% under Xenon and UV irradiation, respectively. Conversely, the introduction of AO had minimal impact on the catalytic performance, decreasing the degradation rate to 97.4% and 98.7% under Xenon and UV irradiation, respectively. These outcomes advise that °OH are the primary active agents in the as-fabricated g-C₃N₄/CQD/Ag material for the photocatalytic decomposition of ACT, with superoxide radicals (°O₂⁻) also playing a role, h⁺ contribute minimally. The findings demonstrate that the as-constructed g-C₃N₄/CQD/Ag generates a substantial amount of °OH radicals under light irradiation, highlighting their notable role in decomposing pollutants. When light is applied to the as-made g-C₃N₄/CQD/Ag material, the CQDs act as electron mediators, accepting and transferring electrons (Equation (2)). This diminishes the reintegration of charge carriers, allowing photo-induced electrons and holes to be efficiently utilized. Electrons are transferred to the Ag via CQDs and subsequently returned to the valence band (VB) by the Ag (Equation (3)). The holes in the VB interact with water to generate °OH, which are crucial for decontaminating persistent organic contaminants (Equations (4) and (5)). Electrons excited to the conduction band can participate in the reduction in dissolved O₂, leading to the formation of °O₂⁻ (Equation (6)). Furthermore, through a sequence of additional reactions (Equations (7)–(13)), °O₂⁻ play an indirect role in facilitating the generation of additional °OH and °HO₂ during the photocatalytic process. Finally, the generated radical species can subsequently interact with ACT, either adsorbed on the surface or present in solution, initiating a series of oxidative reactions (Equation (14)). These reactions result in the formation of various reactive intermediates, ultimately contributing to the degradation of ACT into intermediate and final degradation products [27,40].

$$\mathrm{g}\text{-}{\mathrm{C}}_3{\mathrm{N}}_4/\mathrm{CQD}/\mathrm{Ag}+h\vartheta \to \mathrm{g}\text{-}\mathrm{C}3\mathrm{N}4/\mathrm{CQD}/\mathrm{Ag}(h_{VB}^{+}+e_{CB}^{-})$$

(2)

$$\mathrm{g}\text{-}{\mathrm{C}}_3{\mathrm{N}}_4/\mathrm{CQD}/\mathrm{Ag}(h_{VB}^{+}+e_{CB}^{-})\to \mathrm{Ag}(e_{CB}^{-})+\mathrm{g}\text{-}\mathrm{C}3\mathrm{N}4(h_{VB}^{+})$$

(3)

$$\mathrm{g}\text{-}{\mathrm{C}}_3{\mathrm{N}}_4(h_{VB}^{+})+{\mathrm{H}}_2\mathrm{O}\to °{\mathrm{O}\mathrm{H}}_{\left(\mathrm{abs}\right)}+{\mathrm{H}}^{+}$$

(4)

$$\mathrm{g}\text{-}{\mathrm{C}}_3{\mathrm{N}}_4(h_{VB}^{+})+{\mathrm{OH}}^{-}{}_{\left(\mathrm{abs}\right)}\to °{\mathrm{O}\mathrm{H}}_{\left(\mathrm{abs}\right)}$$

(5)

$$\mathrm{Ag}(e_{CB}^{-})+{\mathrm{O}}_2\to {°\mathrm{O}}_2^{-}$$

(6)

$$\mathrm{Ag}(e_{CB}^{-})+{°\mathrm{O}}_2^{-}+2{\mathrm{H}}^{+} \to {\mathrm{H}}_2{\mathrm{O}}_2$$

(7)

$$\mathrm{Ag}(e_{CB}^{-})+{°\mathrm{O}}_2^{-}+2{\mathrm{H}}^{+} \to {\mathrm{H}}_2{\mathrm{O}}_2$$

(8)

$${°\mathrm{O}}_2^{-}+2{\mathrm{H}}^{+}\to °{\mathrm{HO}}_2$$

(9)

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

(10)

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

(11)

$${°\mathrm{O}}_2^{-}+{\mathrm{H}}_2{\mathrm{O}}_2\to °{\mathrm{O}\mathrm{H}}_{\left(\mathrm{abs}\right)}+{\mathrm{HO}}^{-}+{\mathrm{O}}_2$$

(12)

$$\mathrm{CIP}+(°\mathrm{OH},°{\mathrm{OH}}_2,{°\mathrm{O}}_2^{-}\mathrm{or}{\mathrm{h}}^{+})\to \mathrm{intermediates}\to \mathrm{final}\mathrm{products}$$

(13)

3.2.6. Photocatalytic Decomposition Rate of Previously Published Research on g-C₃N₄ and Its Materials in Comparison with the Current g-C₃N₄/CQD/Ag Material

Table 3. CQD/g-C₃N₄ heterojunctions with different noble metals and their photocatalytic activity efficiency.

Photocatalyst	Illumination Source	Implementation	Degradation Efficiency (%)	Ref
Ag/g-C3N4	200 W tungsten lamps (420 nm cut off filter)	Degradation of MB dye	96	[47]
SDAg-CQD/UCN	350 W Xenon lamps (420 nm cut off filter)	Decontamination of naproxen	87.5	[48]
ZnSe-Ag/CN	300 W Xenon lamps (420 nm cut off filter)	Decontamination of ceftriaxone sodium	89.24	[49]
Ag/g-C3N4	300 W Xenon lamps (400 nm cut off filter)	Decontamination of sulfamethoxazole	87.3	[50]
Ag2O/NCDs@PTCN	300 W Xenon lamps (420 nm cut off filter)	Decontamination of CBZ	99.9	[51]
Ag/CNQDs/g-C3N4	Microwave	Degradation of norfloxacin	100	[52]
CQD/cds/g-C3N4	45 W LED lamps (420 nm cut off filter)	Decontamination of Rhodamine B and sulfadiazine	94.7 & 76	[53]
Ag/UCN	300 W Xenon lamps	Degradation of ofloxacin	95.2	[54]
Ag/UTCN	800 W Xenon lamps (420 nm cut off filter)	Degradation of acetaminophen	85	[55]
Ag/CNQDs/g-C3N4	300 W Xenon lamps (400 nm cut off filter)	Decontamination of norfloxacin, sulfamethoxazole and tetracycline hydrochloride	100 & 81	[56]
g-C3N4/NCDs/Ag	300 W Xenon lamps (400 nm cut off filter)	Decontamination of methyl orange, Rhodamine B, tetracycline hydrochloride and chrysin hydrochloride	100	[57]
Ag/g-C3N4/O3	High pressure Xenon lamp	Degradation of acetaminophen	83.1	[58]
HT-g-C3N4/PS system	LED lamps 400 nm	Degradation of acetaminophen	41	[59]
CQD/CN-PANI	500 W Xenon lamps (420 nm cut off filter)	Decontamination of ciprofloxacin, imidacloprid, tetracycline, phenol and Rhodamine B	94.7, 80.1, 93.2, 45.6 & 97.9	[60]
CDs/g-C3N4	300 W Xenon lamps (400 nm cut off filter)	Decontamination of phenol	87	[61]
Ag/g-C3N4	5 W LED lamps	Degradation of ciprofloxacin	84	[62]
GCN-Ag	200 W Xenon lamps	Degradation of methylene blue, crystal violet and rose bangal	96, 80 & 78	[63]
Br/g-C3N4	38.5W WLEL	Degradation of OTC	75	[64]
Ag/g-C3N4	Solar light	Degradation of tetracycline	96.8	[65]
g-C3N4/TiO2	300 W Xenon lamps (400 nm cut off filter)	Degradation of ciprofloxacin	88.1	[66]
g-C3N4/RGO/WO3	500 W Xenon lamps (400 nm cut off filter)	Decontamination of ciprofloxacin	85	[67]
g-C3N4/TiO2/Kaolinite	Xenon lamps (400 nm cut off filter)	Decontamination of ciprofloxacin	92	[68]
CN-Ce (Cerium ш)	300 W Xenon lamps (420 nm cut off filter)	Decontamination of naproxen	69	[69]
CN-Er (Erbium ш)	300 W Xenon lamps (420 nm cut off filter)	Decontamination of naproxen	44	[70]
CN-Sm (samarium ш)	300 W Xenon lamps (420 nm cut off filter)	Decontamination of naproxen	40.1	[71]
NCDs/g-C3N4	350 W Xenon lamps (420 nm cut off filter)	Decontamination of indomethacin	91.5	[72]
Ag/g-C3N4	300 W Xenon lamps (400 nm cut off filter)	Decontamination of diclofenac	54	[73]
Ag/CDs/CNNs	Sun light	Degradation of methyl orange and p-nitrophenol	98.6 & 92	[74]
Ag/g-C3N4	Sun light	Degradation of morphine	91	[75]
Ag/g-C3N4/ZnO	300 W Xenon lamps (400 nm cut off filter)	Decontamination of paracetamol, amoxicillin and cefalexin	85.3, 41.36 & 71.74	[76]
CQD/g-C3N4	250 W Xenon lamps (420 nm cut off filter)	Decontamination of Rhodamine B and tetracycline hydrochloride	95.2 & 78.6	[77]
Ag/g-C3N4-PNP	300 W Xenon lamps (400 nm cut off filter)	Decontamination of sulfamethoxazole	87.3	[78]
CQD/g-C3N4	300 W Xenon lamps	Decontamination of diclofenac	100	[79]
Ag/g-C3N4/PCN	300 W Xenon lamps (420 nm cut off filter)	Decontamination f tetracycline	83	[80]
AgI/g-C3N4	300 W Xenon lamps (400 nm cut off filter)	Decontamination of diclofenac	45	[81]
HCNs/CDs	350 W Xenon lamps (420 nm cut off filter)	Decontamination of indomethacin, naproxen, ciprofloxacin, norfloxacin, diclofenac, bisphenol A and naproxen	4.8, 98.6, 47.2, 41.7, 10.9 & 20.9	[82]
g-C3N4/CQD/Ag	500 W Xenon lamps (420 nm cut off filter) and UV lamp 250 W (λUV = 254 nm)	Decontamination of acetaminophen	Under visible light = 85.3 Under UV irradiation = 87.5	In this study

presents a comparative summary of multiple photocatalysts presented in the literature for decontaminating different contaminants. It involves a comparison of photocatalytic composites, light irradiation sources, and degradation efficiencies for ACT. The table highlights that doping g-C₃N₄ with noble metallic ions significantly enhances photocatalytic effectiveness for degrading organic contaminants in aquatic solutions. This enhancement is primarily because of the noble metallic ions’ capability to capture electrons, which accumulates around these ions, leading to improved separation of photogenerated holes and electrons [41]. Additionally, the incorporation of CDs into the photocatalyst significantly boosts the relocation of excited electrons from the VB to the CB, thereby enhancing the separation of electrons and holes and declining recombination [42]. This results in improved photocatalytic effectiveness. In the current research, efficient ternary g-C₃N₄/CDs/Ag materials were developed, where CDs act as electron mediators. This configuration demonstrated effective photocatalytic effectiveness under both visible and UV irradiation in comparison with binary photocatalysts. Studies by Zhang et al., Hong et al., and Jiang et al. [43,44,45] have shown that the addition of CDs to g-C₃N₄ promotes the decomposition of stubborn and persistent organic pollutants by generating further reactive oxidation agents. Moreover, our previous published study indicated that loading CQDs and adding noble metal nanoparticles (e.g., Au, Ag, and Pt) onto the surfaces of semiconductor significantly enhance photocatalytic effectiveness. This enhancement is ascribed by the expanded light-harvesting capabilities obtained via the potent surface plasmon resonance (SPR) effects of these noble metals [46].

3.2.7. Determination of By-Products and Mineralization

The mineralization efficiency of the media was evaluated using TOC analysis, as illustrated in Figure 10. After 60 min of reaction with the g-C₃N₄/CQD/Ag catalyst, the degradation rates of ACT were 97.7% and 98.2% under Xenon light and UV illumination, respectively. Correspondingly, the TOC removal rates were 85.11% and 86.89%. These results suggest that ACT and its decomposition intermediate compounds may be almost completely mineralized with prolonged exposure to the catalyst under irradiation. Bianchi et al. [83] presented that more than 85% of TOC was mineralized after 400 min of photocatalytic oxidation of ACT. Similarly, Yang et al. [84] found that over 85% of TOC was mineralized after 450 min of acetaminophen photocatalytic oxidation.

3.2.8. Acetaminophen Degradation Pathway

The chemical structures of potential intermediate substances from the photocatalytic decontamination of ACT were recognized adopting GC–MS analysis, as displayed in Figure 11. The degradation pathways of ACT primarily involve hydroxylation due to the abundant formation of °OH, which are the most dominant reactants. Hydroxyl radicals break C–H bonds, replacing carbon with a hydroxyl group, followed by cleavage of the C–N bond to generate 4-aminophenol [85]. Previous research has shown that ACT can be further converted to hydroquinone and benzoquinones [86]. The benzoquinones are then converted to benzaldehyde and subsequently to benzoic acid, which is additional decontaminated into various carboxylic acids, including formic, oxalic, malonic, malic, and acetic acids. These findings align with those reported by De Luna et al., [87]. who found similar degradation products Consistent with previously published research (Cai et al., 2022; Fan et al., 2019; Feng et al., 2018), [88,89,90] the toxic intermediate compounds produced generated over the ACT decomposition were ultimately converted into harmless carboxylic acids.

3.3. Effect of Dark Conditions on ACT Adsorption by g-C₃N₄/CQD/Ag

In order to determine the amount of adsorption of ACT by the C₃N₄/CQD/Ag composite, the adsorption test was performed in dark conditions (without the use of radiation) (Figure 12). According to the results obtained in Figure 12, the concentration of ACT was limited in the adsorption of the catalyst (less than 9%), and practically, the activity of the catalyst was not specifically observed in the conditions without radiation. Therefore, in this study, it was proved that the appropriate method to reduce the concentration of ACT in aqueous solutions is to combine radiation with a catalyst that can remove ACT completely.

3.4. The Effect of UV Radiation on Photolysis of ACT

The photolysis of ACT was performed under the influence of UV radiation alone and without nanocomposite. According to the results (Figure 13), the efficiency of ACT decomposition (48%) reached its maximum in 150 min and after that the reaction reached equilibrium. In previous studies, the rate of photolysis of ACT with UV radiation has been reported close to the present study [91]. According to the results, the photolysis of ACT can occur at a rate of 48% by UV radiation alone, but this removal percentage is lower than the Photocatalytic process. Therefore, the decomposition of ACT is caused by both photolysis and photocatalyst processes.

3.5. Comparison of the Findings of This Study with Those of Other Studies

A comparative analysis of the findings from this study with those reported in the literature is presented in

Table 4. A comparative analysis of ACT degradation efficiency via photocatalytic methods across various studies.

Catalyst	Light Source (Power Values)	pH	Dose of Catalyst (g/L)	Initial Concentration (mg/L)	Irradiation Time (min)	Removal Efficiency (%)	Reference
Ag-ZnO	Tungsten halogen lamp (300 W)	8.5	1	5	12	90.8	[92]
(P,S)-g-C3N4/TiO2 (5%)	Xenon lamp (30 mW/cm2)	5 and 7	1	20	60	100	[34]
Fe2O3-TiO2	halogen lamp (500 W)	11	1.25	30	180	95.8	[93]
WO3/TiO2/SiO2	Xenon lamp (500 W)	9	1.5	5	240	95	[94]
g-C3N4/CQD/Ag	Xenon lamp (500 W)	7	0.6	50	60	85.3	In this study
	UV lamp (250 W)	7	0.6	50	60	87.5

, focusing on the degradation efficiency of acetaminophen using various nanocatalysts. The results demonstrate that the ternary nanocatalyst g-C₃N₄/CQD/Ag exhibits superior degradation performance under neutral pH conditions and at higher contaminant concentrations compared to previously reported studies.

3.6. Mechanism Elucidation of g-C₃N₄/CQD/Ag Photocatalyst

3.6.1. Band Structure Analysis of g-C₃N₄

The band positions (valence band (VB) and conduction band (CB)) of g-C₃N₄ were calculated using Mulliken electronegativity theory (Equations (14) and (15)) [95]:

$${\mathrm{E}}_{\mathrm{C}\mathrm{B}}=\mathrm{X}-{\mathrm{E}}_{\mathrm{e}}-{\displaystyle \frac{1}{2}}{\mathrm{E}}_{\mathrm{g}\mathrm{a}\mathrm{p}}$$

(14)

$${\mathrm{E}}_{\mathrm{V}\mathrm{B}}={\mathrm{E}}_{\mathrm{C}\mathrm{B}}+{\mathrm{E}}_{\mathrm{g}\mathrm{a}\mathrm{p}}$$

(15)

Here, E_VB and E_CB represent the VB and CB edge potentials, E_gap signifies the band gap of the nanocatalyst, and E_e is the energy difference between the vacuum level and normal hydrogen electrode (NHE) (4.5 eV). The band gap of g-C₃N₄ is estimated to be 2.76 eV (Figure S1). χ is the geometric mean of the electronegativity of the photocatalyst constituent atoms, which is calculated for each semiconductor using the following relation (Equation (16)) [96]:

$$\mathsf{\chi }=\sqrt[\mathrm{i}+\mathrm{i}\mathrm{i}+\mathrm{i}\mathrm{i}\mathrm{i}]{{{(\mathrm{X}}_{\mathrm{a}})}^{\mathrm{i}}{{(\mathrm{X}}_{\mathrm{b}})}^{\mathrm{i}\mathrm{i}}{{(\mathrm{X}}_{\mathrm{c}})}^{\mathrm{i}\mathrm{i}\mathrm{i}}}$$

(16)

In which i, ii and iii are the number of each constituent atoms in the semiconductors. Note that the Mulliken electronegativity (X) for an element is calculated using its ionization energy (E_io) and electron affinity (E_af):

$$\mathrm{X}={\displaystyle \frac{{\mathrm{E}}_{\mathrm{i}\mathrm{o}}+{\mathrm{E}}_{\mathrm{a}\mathrm{f}}}{2}}$$

(17)

Based on Equations (16) and (17), the χ value for g-C₃N₄ was distinguished to be 6.91. By determining the χ value, the E_VB and E_CB of g-C₃N₄ were calculated to be 3.79 and 1.03 V (vs. NHE).

3.6.2. Charge Transfer Mechanism

Under UV and Xenon-light irradiation, g-C₃N₄ absorbs photons, exciting electrons (e⁻) from the VB to the CB, leaving holes (h⁺) in the VB. The ternary composite (g-C₃N₄/CQD/Ag) enhances charge separation and migration:

• CQDs act as electron acceptors due to their lower CB potential. Electrons transfer from C₃N₄’s CB to CQDs’ CB, thermodynamically favoring O₂ reduction [97].
• Ag nanoparticles form a Schottky junction with g-C₃N₄, acting as electron sinks. Electrons migrate from g-C₃N₄’s CB to Ag, suppressing recombination [98,99,100].

3.6.3. Redox Reactions and Radical Formation

As depicted in Figure 14, the E_VB of g-C₃N₄ (3.79 V vs. NHE) is more positive than the redox potentials of OH⁻/HO^• (1.99 V vs. NHE) and H₂O/HO^• (2.34 V vs. NHE) [101]. Therefore, holes accumulating in the valence band of g-C₃N₄ are capable of oxidizing the adsorbed water molecules or hydroxide ions to produce HO^• species:

$${\mathrm{O}\mathrm{H}}^{-}+{\mathrm{h}}_{\mathrm{V}\mathrm{B}}^{+}\to {\mathrm{H}\mathrm{O}}^{•}$$

(18)

$${\mathrm{H}}_2\mathrm{O}+{\mathrm{h}}_{\mathrm{V}\mathrm{B}}^{+}\to {\mathrm{H}\mathrm{O}}^{•}+{\mathrm{H}}^{+}$$

(19)

Furthermore, electrons in the conduction band of Ag and CQD reduce O₂ to superoxide radicals (O₂^•−):

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

(20)

In conclusion, the g-C₃N₄/CQD/Ag ternary composite achieves superior photocatalytic activity by optimizing charge separation and leveraging the high oxidative capacity of HO^•. This mechanism aligns with experimental results showing 85–87% acetaminophen degradation under UV and Xenon light.

3.7. Comparison of the Photocatalytic Performance of g-C₃N₄, CQD, Ag, and Their Composites, in the Photocatalytic Removal Efficiency of Acetaminophen Under UV Light

Figure 15 presents the removal efficiency (%) of acetaminophen under UV light in optimum conditions (C₀ = 50 mg/L, pH = 7, composite dosage = 0.6 g/L and contact time: 60 min), by g-C₃N₄/CQD/Ag, g-C₃N₄/CQD, g-C₃N₄/Ag, CQD/Ag, CQD, g-C₃N₄ and Ag. As shown in this figure, the g-C₃N₄/CQD/Ag composite demonstrated the highest efficiency at 87%, followed by CQD/Ag (76.98%) and g-C₃N₄/Ag (69.87%). Also, the g-C₃N₄/CQD composite achieved 62.27%, while individual materials exhibited lower efficiencies: g-C₃N₄ (26.06%), CQD (39.1%), and Ag (52.07%). Those data highlight the significant enhancement in removal efficiency achieved by the ternary composite compared to its components.

This enhanced performance is attributed to the synergistic effect of the composite materials. g-C₃N₄ provides a large surface area and photocatalytic activity, CQD improves light absorption and charge separation, and Ag nanoparticles enhance conductivity and facilitate electron transfer. The superior efficiency of the composites underscores the improved adsorption and catalytic properties resulting from the interaction between these components. The low efficiency of g-C₃N₄ alone emphasizes the importance of composite design for optimizing removal efficiency.

4. Conclusions

Summarily, efficient g-C₃N₄/CQD/Ag materials were successfully constructed by depositing CQDs and Ag onto the surface of g-C₃N₄ to enhance ACT degradation. The g-C₃N₄/CQD/Ag composites demonstrated improved photocatalytic effectiveness under both Xenon and UV illumination. Various factors, including pH values, photocatalyst concentration, acetaminophen concentration, and reaction time, were investigated. The highest degradation efficiency was achieved with 0.6 g/L of the catalyst, an initial ACT content of 50 mg/L, and a pH of approximately 7. Under these conditions, ACT was effectively degraded within 60 min, with mineralization rates of 97.7% and 98.2% under Xenon and UV illumination, respectively. The decomposition of ACT was primarily driven by °OH. Principal intermediate compounds were detected, and potential degradation pathways were presented. These results indicate that °OH radicals can decompose a wide range of pharmaceuticals into less toxic organic compounds, ultimately releasing low molecular weight products like H₂O and CO₂. Comparing the catalytic rate of the synthesized g-C₃N₄/CQD/Ag materials with other photocatalysts highlights the superior or comparable performance of these composites. g-C₃N₄, when combined with CQDs and Ag, benefits from a narrowed bandgap, effective electron–hole pair separation, and boosted photocatalytic effectiveness, making it an effective photocatalyst for the degradation of ACT.