Abstract
Plastic pollution of the environment and water is a persistent global concern. Microplastic contamination has impacted aquatic environments, demanding the development of efficient remediation techniques to address this issue. In this work, N-TiO2–supported corncob activated (ZnCl2) carbon (CCAC/N-TiO2) composites (CT13) (CT11), and (CT31) were synthesized through a simple wet-impregnation approach. The prepared composite materials were characterized using Fourier transform infrared (FTIR) spectroscopy, XRD (X-ray diffraction), SEM (Scanning electron microscopy), EDS (Electron dispersive spectroscopy), XPS (X-ray photoelectron spectroscopy), thermogravimetric analysis (TGA), Photoluminescence (PL) spectroscopy, UV-visible spectroscopy and Dynamic Light Scattering (DLS) techniques. Tauc’s method was applied to evaluate the optical band gap, and the results indicated that the composite material’s spectral response extended into the visible-light region, accompanied by a significant reduction in band gap energy. The removal performance of the CCAC/N-TiO2 composites toward PVC-NPs was systematically evaluated under different pH conditions (4, 7, and 10), varying contact times, and various light conditions. CT13 composite demonstrated exceptional performance, achieving a 94% degradation efficiency after 180 min of exposure to tungsten light. The removal of PVC-NPs was determined to occur via a photocatalytic pathway and was confirmed by quenching experiments. Additionally, SEM, FTIR, DLS, and fluorescence microscopy verified the presence of PVC-NPs on the composite surfaces under both dark and light conditions. The major photodegradation products were identified using gas chromatography-mass spectrometry (GC-MS). The addition of CCAC to N-TiO2 significantly improved its ability to remove PVC-NPs. This is because the CCAC addition increased the number of active sites for adsorption. The CT13 composite’s surface attracts and captures the PVC-NPs through a variety of interactions, including hydrophobic interactions, electrostatic attractions, π-π interactions, halogen bonding, and hydrogen bonding. This strong adsorption increases the number of available reaction sites, which in turn boosts the photocatalytic removal of the PVC-NPs. This study sheds light on the use of biomass-derived materials for water purification, providing a long-term solution to pollution and agricultural waste issues.
Highlights
This study demonstrates a sustainable and effective approach for the remediation of microplastic contamination—an escalating environmental threat to aquatic ecosystems and drinking water supplies.
Utilization of corncob biomass for preparation of N-TiO2 photocatalysts composite, promotes circular economy and leads to valorization of agro-waste.
The integration of green synthesis with enhanced photocatalytic performance of nanocomposites with degradation efficiency of PVC nanoparticles (94% in 180 min) under light indicates strong potential for practical application in advanced wastewater treatment systems.
This approach offers a scalable and environmentally friendly pathway for mitigating the dual challenge of plastic waste and agricultural residue disposal.
1 Introduction
Plastic pollution is a global issue that could endanger human health and present a serious threat to the ecosystem [1, 2]. Microplastics (MPs) and nanoplastics (NPs) particles are regarded as “one of the most important environmental threats to marine ecosystems” [3]. Particles of micro- and nanoplastics penetrate biological membranes, enter the body via the respiratory system, and bioaccumulate in organs. Therefore, pollution from micro-nanoplastic particles is “a potential threat to food security, health, and environment” [1, 4–6]. The hydrophobic nature of plastic waste also causes contaminants to be adsorbed on its surface, serving as transporters of toxic chemicals and increasing the bioavailability of pollutants to soil as well as aquatic life [7]. Drinking and raw water have been shown to contain microplastics, including polystyrene (PS), polypropylene (PP), polyvinyl chloride (PVC), polyethylene (PE), and others [8–11].
Numerous physical, chemical, and biological techniques have been developed to date to remove microplastics and nanoplastics from water, including coagulation, membrane technology, filtration, ozonation, chlorination, hydrolysis, green algae, active sludges, flocculation and sedimentation, microorganism aggregation, and adsorption. Although the hydrophobicity and surface charge of MPs/NPs facilitate their removal, these methods are often limited by high energy consumption, variable efficiencies, restricted applicability, and potential environmental impacts [12, 13]. Therefore, the development of sustainable technologies capable of degrading MPs/NPs into non-toxic or reusable products remains a critical research priority.
Photocatalysis has become a promising technique for mitigation of plastic pollution due to its high removal efficiency, environmentally benign and cost effectiveness [14, 15]. Among photocatalysts, TiO2 is effective for treatment of toxic wastewater and non-biodegradable organics [16–19]. However, there are certain drawbacks, such as its difficulties in recovery from aqueous and the aggregation of photocatalyst [20]. To address these limitations, TiO2 is often immobilized on high-surface-area supports. Activated carbons are particularly attractive due to their high surface area, porous structure, and ability to enhance charge separation and interfacial electron transfer, thereby improving photocatalytic performance [21–24]. The development of an alternate strategy using TiO2 placed on a support with a large surface area has thereby received a lot of interest.
Corn cobs are an agro-industrial by-product, that is produced in significant quantities during the processing of corn, these are a low-cost and sustainable precursor for biochar production [25, 26, 79] As a low-ash, lignocellulosic carbon-rich material, the cob has been investigated as a precursor for the production of biochar [25, 26]. Corn cob biochar has a significant capacity for adsorption of different organic contaminants [23, 27]. The biochar experienced pore formation as a result of the ZnCl2 activation during the pyrolysis, resulting in increased adsorption sites and a high specific surface and micro porosity [28, 29].
Polyvinyl chloride (PVC) was chosen as the representative MP. PVC is widely used in building materials, daily necessities, artificial leathers, pipes, etc. The main objective of this study is to use nanocomposite of N-TiO2 and carbon derived from corncob to remove polyvinyl chloride from water. In this study, ZnCl2-activated carbon, derived from corncob (CCAC) was adopted as a support for N-TiO2 to produce CCAC/N-TiO2 composites due to its advantages including high porosity, high surface area, high photostability and suitability to work at ambient temperature. The removal of PVC nanoparticles via adsorption and photocatalysis was studied in relation to weight ratio alterations in composite materials. The possible photocatalytic mechanism was investigated so that composite behaviour and properties could be analysed.
2 Materials and methods
2.1 Materials
Corn cobs were purchased from local market of Apra (Punjab). The following commercially available chemicals and solvents were purchased from Sigma-Aldrich: ZnCl2 (≥98%), perylene-3,4,9,10-tetracarboxylic dianhydride (97%), 1-iodobutane (99%), 1,8-Diazabicyclo (5.4.0) undec-7-ene (DBU) (≥99.0%), 1-butanol (99.9%), N, N-dimethylformamide (DMF) (99.8%), acetone (≥99.5%), tetrahydrofuran (THF) (≥99.9%), sodium dodecyl sulfate (SDS) (≥99%), hexane (≥95%), chloroform (≥99.8%), dichloromethane (DCM) (≥99.8%) and silica gel (60–120 mesh), titanium butoxide (97%), ethanol (≥99.8%), ammonium nitrate (NH4NO3) (≥98%). Pellets of PVC of molecular weight ∼ 120,000) were obtained from Hi-Tech Polymer Products in Ludhiana, Punjab.
2.2 Preparation of corncob activated carbon
Corncob was washed in deionized water and dried. 15 g of corncob particles were thoroughly mixed with 60 mL of a 25% (w/v) ZnCl2 solution (impregnation ratio 1: 1). The resultant mixture was allowed to stand for 22 h, dried for 2 h at 100 °C, and then pyrolyzed at 850 °C in nitrogen atmosphere. The resulting activated carbon was mixed with deionized water, vigorously stirred with 1N HCl to get the pH to about 7, and then filtered and dried at 105 °C. The synthesized corncob based activated carbon was named as ‘CCAC’.
2.3 Preparation of nitrogen-doped titanium oxide (N-TiO2)
The sol-gel approach was used to prepare N-TiO2, as explained in our earlier work [30]. Briefly, a 1:1 solution of water and ethanol was used to dissolve the required amount of ammonium nitrate. The N-TiO2 was obtained by adding titanium butoxide dropwise to the previously prepared mixture in a beaker with 70 mL of ethanol. After 3 hours of stirring, the resulting slurry was allowed to age for a day. It was further dried at 100 °C for 12 h and then calcined at 500 °C for 6 h under nitrogen atmosphere.
2.4 Preparation of corncob activated carbon/N-TiO2 (CCAC/N-TiO2) composites
Appropriate amounts of N-TiO2 and CCAC were separately dispersed in deionized water at neutral pH and sonicated for 10 min. The suspensions were then mixed according to the desired CCAC: N-TiO2 weight ratios (1:3, 1:1, and 3:1) to obtain composites designated as CT13, CT11, and CT31, respectively. The combined suspension was adjusted to a total solid concentration of 5 mg mL-1 and sonicated for an additional 20 min prior to further processing.
2.5 Preparation of fluorescent PVC nanoparticles with PTE encapsulation (PVC-NPs)
Perylene-3,4,9,10 tetracarboxytetrabutylester (PTE), a fluorescent dye, was prepared in the lab using a reported procedure [31]. Fluorescent polymer nanoparticles were uniformly dispersed in water, and the resulting solution was referred to as “PVC-NPs” as reported in our previous work [32].
2.6 Adsorption and photocatalytic experiments for removal of PVC from water
The removal activities of the synthesized materials were studied by removal experiments using aqueous solution of PVC-NPs under different conditions. Tungsten bulbs, room light (ambient laboratory illumination), and three different pH values (4, 7, and 10) were used to test the material’s photocatalytic activity in a 20 mL PVC (2 mg/mL) solution using the catalyst dosage of 80 mg per 20 mL of pollutants. The desired pH values (4, 7, and 10) were adjusted using dilute HCl and NaOH solutions prior to the photocatalytic experiments. A 400 W tungsten light (with a maximum emission intensity in the yellow-red region around 550–750 nm) was then used to illuminate the solutions for 3 h with constant stirring. The 3 mL aliquots were taken at regular intervals (30, 60, 90, 120, 150, and 180 min) and centrifuged to remove the catalyst. These samples were evaluated using a UV-Vis spectrophotometer at 446 nm (λmax) to determine the PVC-NPs removal.
To study the adsorption mechanism, the aforementioned experiment using the synthesized materials to remove PVC-NPs from water under the dark conditions. The following formula (Equation 1) was used to determine the removal efficiency in photocatalytic and adsorption experiments:
Where the initial and final concentrations of PVC-NPs solution were denoted by C0 and Cf, respectively.
2.7 Quenching experiment
Scavengers, including sodium oxalate (a hydroxyl radical quencher), isopropyl alcohol (a hole quencher), and para-benzoquinone (a superoxide radical quencher), were used to study the reactive species involved in the photocatalytic reaction. To evaluate the removal efficiencies after the addition of scavengers, 2 mM of each scavenger was added to a mixture of the synthesized material and PVC solution in a setup similar to that used for photocatalytic activity.
2.8 Calculation of point of zero charge
The point of zero charge (Pzc) for the synthesized materials was performed by following the method reported earlier [30].
2.9 Characterization techniques
The prepared samples were characterized using several techniques. X-ray diffraction experiments were performed at room temperature to assess the powders’ crystal phase composition and crystallite size. X-ray diffraction (XRD) patterns were analyzed at room temperature using a Bruker D-8 Advance X-ray diffractometer with a scan range of 2θ = 10–80◦ to assess purity, size, and crystallinity. Based on the XRD results, Scherrer’s equation was used to calculate the average crystallite sizes of the powders. Images from Field Emission Scanning Electron Microscopy (FE-SEM) were obtained using JEOL JSM-7610F Plus. The energy dispersive spectrometry (EDS) analysis was conducted using the Oxford EDS LN2 free, Au Coater; JEOL Smart Coater. The NEXA Surface analyzer from Thermoscientific has been used to record XPS patterns. Fourier transform infrared spectroscopy (FT-IR) spectra of the materials obtained between 4,000 and 400 cm-1 using a Perkin Elmer Spectrum 2 spectrophotometer. A UV 1800S spectrophotometer was used to quantitatively assess PVC-NPs in water using UV–vis spectroscopy both before and after removal at room temperature. Using Tauc’s method, the band gap energies (E.g.,) of the samples were determined. Plotting (αhv)1/2 or (αhν)2 against hν, the intercept of the straight line was determined for both direct and indirect transitions. The LS6500 fluorescence spectrophotometer from Perkin Elmer was used to measure the photoluminescence (PL) spectra. A thermogravimetric analyzer (Perkin Elmer TGA 4000) was used to test the thermal stability of the produced materials. Size and surface charges of the prepared PVC-NPs were measured using a device called the Malvern Zeta sizer Nano-ZS90. Fluorescent images were taken using a confocal microscope (NIKON, model: A1R). After removal treatment, the Shimadzu QP 2010 Ultra GC-MS Gas Chromatography and Mass Spectrometer was used to characterize the PVC-NPs solution.
3 Results and discussion
3.1 Characterization of polyvinyl chloride nanoparticles (PVC-NPs)
The prepared nanoparticles are characterized using FTIR, dynamic light scattering (DLS), UV-visible spectroscopy, and scanning electron microscopy (SEM), as reported in our previous study [32]. To enable optical tracking of the nanoparticles, the fluorescent PTE dye was encapsulated within the particles during the polymer particle synthesis procedure [33].
The fluorescence and UV-Vis spectrophotometer were used to record the optical characteristics of the synthesized fluorescent PVC-NPs. The UV–Vis spectra of the PTE dye encapsulated in the PVC nanoparticles showed two characteristic peaks at 446 and 470 nm, while the emission spectrum revealed a characteristic peak at 550 nm.
Supplementary Figure S1 shows that the PVC-NPs had spherical shapes and an average particle size of 69.9 nm, as determined by SEM study. In order to determine the hydrodynamic size in water, PVC-NPs were further analyzed using dynamic light scattering (DLS), resulting in a particle size range of 120.9–123.7 nm. DLS values are usually greater than the radii obtained from SEM micrographs because SEM measurements establish the dry size of particles, and DLS values indicate hydrated size.
Additionally, Zetasizer was used to further analyze the surface charges of the PVC-NPs dispersed in water at pH 7.0, and the results indicated negative zeta potentials between −21.5 and −30.3 mV.
Supplementary Figure S2 displays the FT-IR spectra of PVC-NPs and bare perylene-3,4,9,10-tetracarboxytetrabutylester (PTE). PVC-NPs exhibited similar peaks following the addition of PTE into PVC, but some additional PVC-specific peaks were observed.
3.2 Characterization of materials
3.2.1 Scanning electron microscopy (SEM)
The field emission scanning electron microscope (FESEM) was used to analyze the morphology of the synthesized composite materials. The well-organized and regular arrangement of the honeycomb-shaped surface pores on ZnCl2-impregnated corncob activated carbon results in a porous structure [34]. SEM images of CCAC indicated a highly porous morphology with evenly distributed spherical hollows and a wide range of pores. The evaporation of ZnCl2 and the decomposition of water and other organic molecules during the carbonization process resulted in a pore-forming action, which increased the surface area of the activated carbon. The improved adsorption capacity of CCAC may have been facilitated by its rougher and more porous surface [35, 36] shown in Figure 1a.
FIGURE 1
Figures 1b–d displays the FESEM images of the CT13, CT11, and CT31 composites. As demonstrated in our earlier study, the pure N-TiO2 nanoparticles have a spherical crystalline shape [31]. The images of CCAC/N-TiO2 composites in Figure 1 showed that tiny N-TiO2 granules are well dispersed on the CCAC. The addition of CCAC to the composites does not seem to have significantly altered the morphological characteristics of N-TiO2. In CCAC/N-TiO2 composites, N-TiO2 particles gradually accumulated on the outer surface of the CCAC with an increased weight ratio of N-TiO2, leading to pore blockage. The composite also contained some bigger N-TiO2 particles. According to these results, the N-TiO2 particles’ increased free energy in the nano system caused a decrease in relative stability, which in turn caused particle aggregation because of intrinsic instability [37]. As a result, the SEM image showed comparatively some large N-TiO2 particles. Increasing the amount of N-TiO2 dispersed across the CCAC surface may also result in the formation of a sufficient amount of photocatalytic radicals.
3.2.2 Energy dispersive X-ray (EDX) analysis
The elemental composition of all synthesized materials was investigated using energy dispersive X-ray (EDX) analysis. The EDX spectrum and quantitative elemental composition of CCAC, CT13, CT11, and CT31 are shown in Supplementary Figures S3-S6. The findings showed that there was no zinc accumulation on the surface of CCAC, but there was a significant percentage of C and O. The presence of C, O, N, and Ti in CCAC/N-TiO2 composites was verified by elemental mapping, indicating that N-TiO2 nanocrystals have been sufficiently incorporated or distributed on the honeycomb-shaped surface of CCAC. The successful formation of the CT13, CT11, and CT31 composites without impurities was demonstrated by the presence of C, O, Ti, and N peaks. As the ratio of CCAC to N-TiO2 increased, the weight ratio of carbon also increased.
3.2.3 Fourier-transform infrared spectroscopy (FTIR)
The FTIR spectra of bare CCAC, N-TiO2, and the structural changes caused by the interaction of CCAC and N-TiO2 in composites CT13, CT11, and CT31 are shown in Figure 2. For FTIR spectrum of CCAC, the peaks at 3,663 cm-1, 2,978 cm−1, 2,869 cm−1 represents the presence of stretching vibration of hydroxyl (-OH), asymmetric stretching of–CH2 groups and symmetric stretching of–CH3 groups of the organic aliphatic chains, respectively. The vibrations at 1,018 cm−1 shows C-O stretch (due to ethereal group), whereas a sharp band at 1728 cm−1 is allocated to carbonyl functionalities. Aromatic compounds may exhibit a medium intensity band at 1,559 cm−1. Characteristic absorption peaks of lignin were observed at 1,601 cm-1 and 1,567 cm-1. Stretching vibrations of C=C and C-O on the aromatic ring of lignin correspond to 1,280 cm-1, indicating that cellulose, lignin, and hemicellulose are the sample’s major structural components [38]. The bending vibrations of the C-H groups were responsible for the various bands at 1,398 cm−1, while the O-C, C-H, and O-H groups were responsible for the bands at 1,136 cm−1, 773 cm−1, and 604 cm−1, respectively. The presence of C-O groups was demonstrated by the three peaks that appeared at 1,136, 1,060, and 1,009 cm-1. These findings demonstrated that CCAC has a variety of surface-level oxygen-containing functions [38].
FIGURE 2
Figure 2 represents the FTIR analysis that successfully identified the modifications to the surface functional group of the CT13, CT11, and CT31 composites. The stretching vibration of C-H on the aliphatic molecular chain is responsible for the peaks at 2,978 cm-1 and 2,911 cm-1. Ti–OH bending vibrations were represented by the signals at 1,572 cm−1, whereas Ti–O–C stretching vibrations were represented by the signal at 1,069 cm−1. This implies the formation of chemical linkages between CCAC and N-TiO2. The Ti–O–Ti stretching vibration signals at 480–660 cm−1 are part of the distinctive N-TiO2 peaks [47]. The establishment of the oxygen–metal–oxygen bonding was shown by the peak at 759 cm−1 and the narrow peak at 553 cm-1, confirming the formation of activated carbon-loaded TiO2 nanoparticles. The peak at 2,984 cm−1 resulted from the OH-stretching on the surface and reduced their intensity because of the CCAC loading. Stretching vibrations of the C-O bond are indicated by the peak at 1,218 cm−1. The synthesis of N-TiO2 nanoparticles loaded with CCAC was identified by the C-O-Ti-O bonding, which was characterized by peaks at 824 cm−1 and 567 cm−1. Thus, FTIR spectra provide a clear indication of the development of Ti-O-C in CT13, CT11, and CT31 samples, which can alter optical and charge transfer properties [39], demonstrating that N-TiO2 modified the surface structure of CCAC after the synthesis procedure.
3.2.4 X-ray diffraction (XRD)
The crystallinity of pure CCAC, N-TiO2, and CCAC/N-TiO2 (CT13, CT11, and CT31) samples were investigated by X-ray diffraction (XRD) analysis shown in Figure 3. X-ray diffraction (XRD) pattern of N-doped TiO2, confirming the presence of a pure crystalline anatase phase of titanium dioxide (TiO2). The 2θ values at 25.3°, 36.9°, 37.8°, 38.5°, 48.0°, 53.9°, and 55.0° correspond to the specific crystalline planes of (101) (103) (004) (112) (200) (105), and (211), which are characteristic of the tetragonal anatase structure. CCAC’s XRD peaks appeared at 25.7°, 28.2°, 31.8°, 34.6°, 36.2°, 42.8°, 47.5°, 56.5°, 63.1°, and 69.3° indicating the presence of (020) (200) (201) (112) (101) (212) (102) (110) (114), and (340) orientation planes. The broad diffraction features observed in the XRD pattern confirm the predominantly amorphous nature of the activated carbon. The additional sharp diffraction peaks are attributed to residual inorganic mineral phases originating from the corncob precursor, such as potassium-, calcium-, magnesium-, or silica-containing compounds, which remain after carbonization and activation. These peaks represent crystalline inorganic constituents rather than graphitic carbon domains [79]. Characteristic peaks of CCAC and N-TiO2 corresponding to the anatase phase (PDF-00–064–0863) were present in high intensity in the XRD patterns of CT13, CT11, and CT31 composites. This means that the preparation process of CCAC/N-TiO2 composites did not change N-TiO2 crystal phase [40]. Average crystalline size of N-TiO2, CCAC, CT13, CT11 and CT31 were also calculated using XRD analysis through Debye Scherer formula as 13.69 nm, 40.39 nm, 28.15 nm, 28.76 nm and 33.86 nm respectively. Crystallite sizes of CCAC/N-TiO2 composites are found to be greater than N-TiO2. Crystallite sizes of CCAC/N-TiO2 composites decrease as the amount of CCAC decreases from CT31 to CT11 to CT13. A decrease in crystallite size results in increased surface areas, which enhanced reactant adsorption and, in turn, photo-reactivity [41–43]. Greater surface area from smaller crystallites results in more surface-active sites, which raises photonic efficiency. Furthermore, the charge carriers’ diffusion path length from the photogenerated site to the surface site of reaction is shortened by the sample’s lower crystallite size. Photocatalytic efficiency is increased when this diffusion path length is decreased because it lowers the rate at which photogenerated charge carriers recombine [44].
FIGURE 3
3.2.5 Optical band gap
Band gap energy (Eg) can be deduced by extrapolation from Tauc plot shown in Figure 4. It was found that N-TiO2 has a band gap of 2.95 eV, which is lower than that of pure TiO2. In addition to extending the optical absorption of TiO2 to the visible light region and improving the visible light response, nitrogen doping lowers the band gap of TiO2 by introducing a new mid-gap state created by nitrogen incorporation or by mixing N 2p and O 2p states on top of the valence band [45]. Interestingly, the CCAC/N-TiO2 composites CT13, CT11, and CT31 exhibit the band gaps within range of 2.66- 5.1 eV, 2.98–5.1 eV, and 2.80–4.4 eV respectively. The formation of new energy states in the composites due to the development of chemical bonds between the constituents is the cause of this phenomenon [46], which has been seen in intercalated composites previously [47]. The FT-IR and XPS analysis further supports these findings, which imply that Ti–O–C bonds are formed as a result of the N-TiO2 interaction with the CCAC. This could be related to the formation of new heterojunction structures or defects during the synthesis of CT13 [48, 49]. According to reports, carbonaceous compounds can reduce the band gap by acting as impurities through this type of bond [50].
FIGURE 4
3.2.6 Photoluminescence spectroscopy (PL)
Electron-hole pair separation efficiency is commonly measured using photoluminescence (PL) analysis. A greater signal intensity frequently denotes a higher rate of recombination of these charges and, consequently, a lower level of overall photocatalytic activity because photogenerated electrons and holes recombine to produce photoluminescence emission [51–53]. All of the catalysts displayed comparable emission spectra when excited by ultraviolet light at 266 nm, as seen in Figure 5. It was observed that the emission band’s intensity was significantly decreased by the addition of CCAC. CCAC had the potential to prevent recombination of photoelectrons and holes, as indicated by the notable decrease in PL peak intensity. According to reports, there were three primary causes of the decline in PL intensity. (1) Photogenerated electrons could be effectively transported from N-TiO2 to CCAC in the presence of a CCAC/N-TiO2 heterojunction, increasing the electron-hole pair separation efficiency [59]. (2) The quenching effect was brought about by the CCAC’s absorption of photoluminescence [54]. (3) Low charge recombination could result from low electron photoexcitation [55].
FIGURE 5
Reportedly, biochar, activated carbon, and other carbonaceous materials can lower the photoluminescence signal by acting as an acceptor and conductor of electrons, facilitating the transport of photogenerated electrons on the TiO2 surface, and promoting the separation of electron-hole pairs [56–58]. Activated carbons are made up of complex aromatic-aliphatic organic compounds in addition to pores and minerals. Among these aromatic components are compounds with conjugated bonds that are resonant, like graphene, and are what give activated carbons their conductive capacity (electron storage/transport), which makes them appealing for photocatalysis [59]. As the temperature of pyrolysis and carbonization increases, the lignin, cellulose, and hemicellulose molecules in the biomass undergo dehydrogenation, hydrogen transfer, and isomerization reactions, resulting in the formation of these locally bonded aromatic rings. As the temperature rises, the aromatic ring sheets stack into a turbostratic pattern [60]. Conjugated activated carbon structures act as an electron acceptor, providing a pathway for photogenerated electrons from TiO2 to migrate and separate from the holes through the π-conjugated systems. This separation of electrons and holes significantly reduces the rate of recombination, enhancing the overall photocatalytic activity of the composite material [59].
3.2.7 Thermogravimetric analysis (TGA)
The thermal stability of the composites was evaluated by thermogravimetric analysis (TGA) at a heating rate of 10° C/min in an inert atmosphere. The TGA curves for composite materials CT13, CT11, and CT31, as well as bare N-TiO2 and CCAC (inset), are shown in Figure 6. These curves involve three weight loss stages up to 600 °C. The weight loss of CCAC was approximately 78.55% at 600 °C. Since cellulose, hemicellulose, and lignin have distinct chemical compositions and levels of thermal stability, they decompose at various temperatures. The three steps of the pyrolysis process-drying, decomposition, and carbonization, are determined by the TGA curve of corncob. A mass loss of around 19.04% occurred during the first step, or drying stage, between 30 °C and 75 °C, confirming that the corncob sample had lost moisture and some volatile materials. The primary mass loss stage was the second stage, often known as the disintegration stage. At about 210 °C, mass loss began, and it ended at about 600 °C. Inside the reactor, hemicelluloses, cellulose, and a portion of lignin underwent pyrolytic breakdown under this temperature range. The breakdown of cellulose and hemicellulose starts about 220 °C–340 °C and 340° C–400° C, respectively [66]. Lignin is the most difficult to break because it is more thermostable and covers a wide temperature range of up to 800 °C–900 °C. The first range of carbonization is 400° C since the minute degradation that occurs before it due to some non-combustible biomaterial can be disregarded. [61–63]. On the other hand, bare N-TiO2 sample exhibits stability up to 600 °C, with a total weight loss of 1.46 weight %. Its remarkable thermal stability is most likely the result of calcining it for 6 hours at 500 °C. The loss of residual carbon from the gum matrix may have contributed to the slight linear reductions in weight that were seen in the 100 °C–600 °C range [64]. The weight loss curve for the CCAC/N-TiO2 composite was composed of three steps. The first thermal event of CCAC/N-TiO2 weight loss occurred between 30 °C and 220 °C due to the loss of physiosorbed water and certain organic solvents adsorbed on the composites’ surface [65]. Next, the subsequent weight loss was attributed to the loss of hydroxyl groups on N-TiO2 and the degradation of cellulose and hemicellulose, this weight loss became more noticeable at temperatures between 220 °C and 340 °C [66]. Finally, the third weight loss was likely due to the exothermic process - which involves the burning of carbon - was the cause of yet another notable weight loss at temperatures higher than 340 °C [67, 68]. The incorporation of N-TiO2 placed on the CCAC substrate surface reduced the weight loss of CCAC/N-TiO2 composites. Briefly, it is suggested that the composite’s thermal stability might be improved by adding N-TiO2 nanoparticles to CCAC. As the loading of N-TiO2 deposited on the surface of the CCAC substrate increased, the weight loss of the CCAC/N-TiO2 composites reduced. This is because the main weight loss in the TGA curves of these composites is attributed to the decomposition of the activated carbon (AC) phase, and as the N-TiO2 concentration increases, the amount of AC decreases, leading to a lower overall weight loss.
FIGURE 6
3.2.8 X-ray photoelectron spectroscopy (XPS)
XPS was used to assess the chemical composition and valency of the synthesized CCAC/N-TiO2 composites. To provide further support for the successful deposition of N-TiO2 onto the CCAC surface, a thorough XPS spectra of the CCAC/N-TiO2 composites CT13, CT11, and CT31 are shown in Figure 7. X-ray photoelectron spectra investigation (Figure S-7a) revealed that all three CCAC/N-TiO2 composites CT13, CT11, and CT31 were made up of carbon, nitrogen, titanium, and oxygen. For Ti 2p, O 1s, N 1s, and C 1s, the corresponding binding energies are 458.6, 531, 400, and 283.7 eV.
FIGURE 7
Ti 2p3/2 and Ti 2p1/2 were linked to two peaks in the Ti 2p spectra, located at 459.4 eV and 465.4 eV, respectively [32]. The Ti 2p peaks were in good agreement with pure N-TiO2. The results show that the titanium signal in the samples increases as the N-TiO2 ratio in the impregnation increases in CT13.
The O 1s core level spectra were in agreement with FTIR data and fitted to two Gaussian curves with centers at 530.9 eV and 532.8 eV, which correlated to -OH on the surface of activated carbon and Ti-O-Ti bonds. Figures 7e–g shows a peak in the O 1s spectrum at 239 eV, indicating the existence of Ti-O. Furthermore, O 1s have a lattice oxygen spectrum at 530 and 531 eV, and their association with O-C=O and C=O, respectively is due to activated carbon.
The deconvoluted C1s spectrum of the CT13, CT11, and CT31 composites is shown in Figures 7i–k, respectively, where the signals for Ti-C, C-C, C=C, C-O, Ti-O-C and, π-π are located at ∼ 284.7, ∼285.5, ∼286.1, ∼ 287, ∼ 289.4, and ∼ 293.9 eV, respectively [69]. The relative intensity of the C-C signal indicated that the CCAC in the CCAC/N-TiO2 composites had a graphite-like carbon structure [70]. The XPS data further verified the structural composition of the CCAC/N-TiO2 composites and confirmed the earlier reported FTIR results. Figures 7i–k shows that the CT13 sample had a larger ratio of C-C (hydrophobic) bonds to hydrophilic bonds (C-O, C = O, -COO) compared to the other two samples. In CT13, the peak area of C=C declines while the peak area of C-C increases. This indicates a conversion of C=C sp2 hybridization (graphite state) to C-C sp3 hybridization. Instead of just intermolecular binding, it is demonstrated that the heteroatoms (such as O, Ti, etc.) are linked on CCAC by chemical bonds [71]. The incorporation of these surface functionalities may alter the surface wettability and moisture interaction behavior of the composites. In photocatalytic systems, hydrophilic surface groups can interact with adsorbed water molecules to generate hydroxyl radicals (•OH), which are highly reactive oxidizing species involved in the degradation of organic pollutants [72]. Optimizing the synthesis conditions of CCAC/N-TiO2 composites might improve their surface wettability and pollutant–surface interactions, resulting in greater interactions with pollutants, which may be useful in their removal [24, 72].
3.3 Removal of PVC-NPs by using synthesized composites under different conditions
A number of experiments were carried out to determine the removal efficiency of the composite material (Figure 8) under dark, tungsten light and room light conditions. It was observed that in dark, PVC removal performance is greatly influenced by the surface area of the material, because it is expected that adsorption dominated the removal of PVC-NPs under dark conditions. CCAC and pure N-TiO2 achieved the highest PVC-NPs removal of 33.58% and 24.98% followed by the CCAC/TiO2 composites (48.6%, 35.6% and 31.7% for CT13, CT11 and CT31 respectively) as shown in Figure 8a.
FIGURE 8
The efficiency of composites in removing PVC-NPs was subsequently examined in both tungsten light (Figure 8c), and room light (Figure 8b). Under these conditions, CT13, CT11, and CT31 demonstrated greater removal of PVC-NPs than under dark conditions; CT13 reached the highest percentage removal of 94% in tungsten light. It could be said that photocatalytic degradation would also concurrently occur with adsorption as light interacts with the material. Additionally, the percentage removal of PVC-NPs by pure N-TiO2 under light conditions improved from 35.7% (dark) to 47.5% (room light) and 59.9% (tungsten-light), demonstrating the photocatalytic effects of N-TiO2.
After 180 min, the CCAC/TiO2 composites had higher PVC-NP removal efficiencies than bare CCAC, owing to the synergistic effect of adsorption properties characteristics of CCAC and the photocatalytic properties of N-TiO2. PVC-NPs were efficiently adsorbed by the porous structure of the CCAC, and their photocatalytic destruction was aided by the N-TiO2. The composite material and PVC-NPs came into contact as a result of this combination, which made removal comparatively effective.
The degradation rates of PVC-NPs for three synthetic composites are displayed in Figure 8. The CT13 composite with a CCAC: N-TiO2 ratio of 1:3 showed the highest removal rate. The CCAC/N-TiO2 composites exhibit a photocatalytic activity order of CT13 > CT11 > CT31. These findings suggest that maximum photocatalytic efficiency can be achieved at a specific N-TiO2 loading amount. The removal rate, however, started decreasing when the CCAC loading increased over this point. A real-life illustration of PVC-NPs degradation by CT13 composite material at pH 4 is shown in Figure 8b, showing a transparent and clear liquid. The photoactivity of N-TiO2 was suppressed by the catalyst with the highest CCAC concentration, because it hinders photon uptake by the N-TiO2 surface [73]. The well-dispersed N-TiO2 particles on the surface of CCAC, as confirmed by XRD pattern (Figure 3), may be responsible for the improvement of the photocatalytic rates for the CCAC/TiO2 composites up to a ratio of 1: 3 (in CT13) as compared to neat N-TiO2, because the PVC-NPs that are adsorbed on the CCAC surface can be readily transferred to the N-TiO2 surface.
3.3.1 Effect of pH
For the removal of PVC-NPs using prepared CT13, CT11 and CT31 nanocomposite at pH 4, pH 7 and pH 10 have been studied. The photodegradation process can be optimized by evaluating the catalyst’s surface charge properties in relation to the pH of the water. The point zero charge (PZC) of the N-TiO2 is 5, indicating that at pH values lower than this, the surface charge of N-TiO2 is positive (pH < PZC), on the other hand, the surface charge turns negative as the pH rises above 5 [74]. However, introducing CCAC changed the PZC value, moving it from acidic to neutral pH [75]. The PZC values for CT13, CT11, and CT31 were 6, 6.15, and 6.37, respectively. The pH value increased with an increasing percentage of CCAC in CCAC/N-TiO2 composites. According to the results, CCAC/N-TiO2 composites showed the maximum removal capacity of 94%, 92.7%, and 51.4% under tungsten bulb light for CT13, CT11, and CT31 composites, respectively, for PVC-NPs (having a negative surface charge, as discussed in Section 3.1) in an acidic medium at pH 4, as shown in Supplementary Table S1, because the surface of CCAC/N-TiO2 composites was positively charged (pH < PZC) and attracts negatively charged pollutants when the pH is below point zero charge (PZC) [76]. This indicates that adsorption is dependent on electrostatic interactions [77]. By increasing the pH from 4 to 7, CCAC/N-TiO2 gradually lost its positive charge until it reached the PZC of composites, i.e., ∼6, and the adsorption capacity of PVC-NPs decreased. At higher pH levels (7 and 10), both PVC-NPs molecules and CCAC/N-TiO2 composites (pH > PZC) were negatively charged, resulting in a significant decrease (Supplementary Table S1) in adsorption capacity due to increased electrostatic repulsions and, ultimately, photocatalytic activity [78].
3.3.2 Effect of light source
Furthermore, it was observed that the removal efficiency was greatly influenced by changing the light sources and at different pH as shown in Supplementary Table S1. Tungsten bulb and room light have been used in a number of experiments. The visible light spectrum of room light ranged from about 400 nm to 780 nm [79], whereas the wavelengths of tungsten bulb light were between 350 and 1,000 nm [80]. According to the experimental results, the rate of degradation under room light was less than that under exposure of tungsten bulbs. This is due to the fact that the wavelength of the light source significantly impacts a photocatalyst’s ability to absorb light and subsequently drives a reaction. Shorter wavelengths (higher energy) range are generally more effective for exciting electrons across the band gap of photocatalysts, leading to higher photocatalytic activity [81].
3.3.3 Effect of scavengers
The common substances involved in photocatalytic reactions that contribute to the degradation of organic matter are holes, hydroxyl radicals, and superoxide anion radicals [82]. To investigate the degradation mechanism of CCAC/N-TiO2 composites on PVC-NPs, active species trapping experiments were conducted. Isopropyl alcohol, sodium oxalate, and parabenzoquinone were used to quench hydroxyl radicals, holes, and superoxide anion radicals, respectively. The addition of isopropyl alcohol, sodium oxalate, and parabenzoquinone decreased the removal efficiency of PVC-NPs to 6.8%, 22.1% and 20.6%, respectively, as illustrated in Figure 8d, suggesting that the breakdown of PVC-NPs was inhibited to different extents. As a result, it was determined that all three active species were involved in the oxidation of PVC-NPs, with hydroxyl radicals playing a major part. Zhao et al. reported similar results [83].
3.3.4 Effect of time
Another important aspect was how removal time affected the synthetic composites’ capacity to extract PVC-NPs from water shown in Figure 9. PVC solutions were agitated with composites at intervals of 30–180 min in order to examine their removal efficacy. The elimination rate of PVC-NPs was assessed using a photocatalyst dosage of 80 mg (based on preliminary trials). As the duration increased from 30 to 180 min, the removal rate (%) of PVC particles for CCAC, N-TiO2, CT13, CT11, and CT31 at pH 4 under tungsten bulb light increased from 18.4% to 38.7%, 17.8%–59.9%, 36%–94%, 24.6 %to 92.7%, 21.3%–51.4%, respectively (Supplementary Figure S8). The catalyst’s surface receives more light energy over time, which causes photoexcited species to develop and increases photocatalytic activity. Real-life image of PVC-NP solution in the time interval of 30–180min at pH 4 is shown in Figure 9b.
FIGURE 9
3.4 Analysis of CT13 and PVC solution after removal experiment
A further study was carried out on CT13 and the solution left behind after removal experiment to better understand the mechanism involved in the process. After the process, a real-life image of PVC-NPs adsorbed on CT13 composite at pH 4 was clearly visible and decreased at pH 7 and pH 10. A visible color change as well as swelling of CT13 at pH 4 was observed after the reaction, as shown in Figure 10. The swelling of CT13 composite could be attributed to adsorption-induced swelling, which refers to the expansion or deformation of a porous material when it adsorbs a pollutant [84].
FIGURE 10
Figures 11a,b display SEM micrographs of the surface morphology of CT13 composite with and without light, respectively, after the removal of PVC-NPs. The findings revealed that there are some distinct spherical particles of PVC-NPs on the surface of CT13, but the distinction is difficult because both N-TiO2 and PVC-NPs exhibit spherical surface morphology. As shown in Figures 11a,b, elemental mapping and EDS analysis verified the presence of chlorine, indicating that PVC-NPs interacted with the CT13 composite after the removal reaction.
FIGURE 11
Additionally, after the removal experiment, the presence of yellow-dyed PVC nanoparticles on CT13 composite was examined using a fluorescence microscope. Fluorescence microscopic images of the left over CT13, CT11, and pure PVC-NPs after removal experiment are displayed in Figure 12. The images demonstrated that after interacting with PVC-NPs, green, fluorescent PVC nanoparticles remained on the surfaces of CT13 and CT11, suggesting the existence of significant interactions between composites and PVC-NPs.
FIGURE 12
In order to evaluate changes in particle zeta potential, DLS measurements were performed on solution obtained right after the removal experiment. After the removal of the PVC nanoparticles, the observed zeta potential of the solution was −13.9 mV. The values are lower than those of pure PVC-NPs, possibly due to charge neutralization by the CT13 composite [85].
The composition of the CT13 composite and pure PVC solution before and after removal experiment was investigated using FTIR spectroscopy, as seen in Figure 13a. The FTIR spectrum of PVC displays peaks at 2,965 cm−1 and 2,924 cm−1, indicating the CH2 asymmetric stretching vibration mode. The peaks at 1,438 cm−1 and 1,282 cm−1 correspond to the C–H aliphatic bending bond and the C–H bending bond near Cl, respectively. The PVC backbone chain’s C-C stretching bond is found between 1,000 and 1,100 cm−1. Lastly, the C-Cl bond is represented by peaks in the range of 600–650 cm−1. After adsorption, FTIR spectra of CT13 displays the addition peaks corresponding to PVC NPs, which point to the existence of PVC NPs adsorbed on CT13 surface. After photocatalytic activity, FTIR spectra of CT11, displays the addition peaks in 850–515 cm-1 region, which corresponds to presence of C-Cl bond [86]. Supplementary Figure S9 represents the FTIR spectra of PVC solution before and after treatment.
FIGURE 13
3.4.1 Intermediates and degradation pathway
Degradation products of PVC-NPs must be identified in order to understand the mineralization process and its effects on the environment. The intermediates during the oxidation of PVC-NPs were assessed using Gas Chromatography-Mass Spectrometry (GC-MS). A potential degradation mechanism of PVC-NPs was suggested in Scheme 1 based on the GC-MS analysis of short-chain organics. In addition to releasing organic intermediates into solution, PVC-NPs were oxidized by •OH to produce oxyorganics with oxygen groups like C=O and O-H. The exfoliating hydrocarbon species were then subjected to further oxidation by •OH to break down into lower molecular weight components like alcohols, carboxylic acids, and esters. These materials eventually partially mineralized into CO2 and H2O. The mass fragmentation pattern of degraded products was used to reveal their structures after correlating with the reported literature [87].
SCHEME 1
The vapours produced from the photocatalytic reaction flask was passed through lime water, which turned milky and indicated that vapours might be containing CO2. Furthermore, the mineralization efficiency of PVC-NPs solution with GT13 at pH 4 was investigated using Total Organic Carbon (TOC) analysis. The reduction of 96.9% in TOC indicated the transformation of PVC-NPs into mineralized products.
3.4.2 Mechanism of removal of PVC-NPs
As seen in Scheme 2, the composite’s removal mechanism was ascribed to the synergistic interaction of CCAC adsorption and N-TiO2 photocatalysis. Firstly, PVC-NPs are adsorbed by CCAC by hydrophobic interactions, π-π interaction, electrostatic attractions, halogen bonding, and hydrogen bonding between CCAC and PVC-NPs (Scheme 2) [78, 88–97]. The π-electrons from the benzene rings of CCAC and the halogen atoms of PVC form halogen bonding. While the benzene rings and hydroxyl groups (-OH) can be thought of as electron-donating groups, the -Cl on the PVC branch may operate as an electron acceptor. This could result in the formation of a particular halogen bond between the -Cl groups and electron donor arenes [97]. Aromatic rings of CCAC and polyene structures in PVC-NPs show π-π interactions, enhanced sorption [98]. PVC-NPs occupy active adsorption sites on the hydrophobic surface of CCAC/N-TiO2. The experimental results indicate the presence of hydrophobic interactions between PVC plastic surfaces and the hydrophobic surface of CT13, as confirmed by XPS (Section 3.2). This is consistent with current study findings, which show that synthetically generated [sorbents are highly effective at adsorbing PVC-NPs from aqueous bodies using hydrophobic interactions as the primary sorption mechanism [5]. CCAC can lower the rate at which electrons and holes recombine, increasing the quantity of electrons accessible for photocatalytic processes. Additionally, by dispersing N-TiO2 particles more widely, CCAC can reveal additional photocatalytic reaction active sites [70].
SCHEME 2
CCAC enhances the generation of •OH radicals and positive hole formation in N-TiO2. This is because electrons are transferred from the conduction band of N-TiO2 particles to the graphitic structure of CCAC, which lowers electron-hole recombination. This is in good agreement with a recent study by Mao’s group on the use of TiO2/carbon black composite catalysts for the degradation of organic contaminants in the liquid phase. Additionally, they noted a comparable improvement in the photocatalytic activity of TiO2 combined with carbon black [72, 99, 104, 105].
It is anticipated that PVC-NPs degradation using CCAC/N-TiO2 samples will proceed as follows, according to a well-established photocatalysis procedure. (i) Light radiation creates electron-hole pairs in N-TiO2 nanoparticles in CCAC/N-TiO2 composite (Equation 2). Adsorbed and/or dissolved oxygen interact with photo-excited electrons in the N-TiO2 conduction band to form superoxide free radicals (•O2¯) (Equation 4), and (ii) passage from the N-TiO2 conduction band to the CCAC domains (Equation 3), where it accumulates and reacts with adsorbed or dissolved oxygen to form superoxide free radicals (•O2¯) (Equation 5) [100]. (iii) N-TiO2 valence band holes react with hydroxyl ions in aqueous PVC-NP solution, forming hydroxyl free radicals (•OH) (Equation 6). (iv) These radicals are highly reactive with PVC-NP molecules, degrading to less toxic molecules such as CO2, water and other degraded products (Equations 7, 8) [101, 102]. The use of CCAC as a PVC-NP adsorbent improves the interaction between N-TiO2 nanoparticles and PVC-NPs throughout the process. The CCAC substrates in the CCAC/N-TiO2 provide an effective electron channel via Ti-O-C bonds for photo-excited electrons in the N-TiO2 conduction band to flow towards the CCAC surface, where they react with adsorbed/dissolved O2 to generate PVC-NPs reactive superoxide free radicals (•O2¯) [77, 103].
4 Conclusion
In summary, ZnCl2 activated carbon derived from corncob biomass (CCAC) was used as a support material for N-TiO2 nanoparticles. The heterogeneous photocatalytic material composed of CCAC and N-TiO2 nanoparticles altered the N-TiO2 lattice system, assisting in lattice distortions and bandgap reduction, improving surface morphology, and forming C-Ti-O bond. The CCAC entrance boosted visible light absorption and produced superoxide ions and hydroxyl radicals on the surfaces. Reduced electron-hole pairs and carrier migrations improved photocatalytic activity against PVC-NPs in CT13 (94%) compared to N-TiO2 nanoparticles (59.9%). Overall, the results suggest that the efficiency of catalysts for varied CCAC/N-TiO2 ratios is determined by the synergistic action of adsorption and photocatalysis. The presence of CCAC enhances PVC-NP adsorption and diffusion from CCAC to N-TiO2. The CCAC matrix reduces electron-hole pair recombination rates in N-TiO2, CCAC acts as an electron sink, trapping photogenerated electrons from N-TiO2 and preventing them from recombining with photogenerated holes and enhanced electron-hole separation, leading to increased •OH production in solution, which significantly improves photocatalytic activity. CCAC/TiO2 composite catalyst is inexpensive, non-toxic, and has a good photocatalytic efficacy, it would be regarded as a new, cost-effective, and environmentally friendly photocatalyst for eliminating pollutants from the aqueous phase. Consequently, the CCAC/TiO2 composite may find use in applications related to wastewater remediation development.
Statements
Data availability statement
The original contributions presented in the study are included in the article/Supplementary Material, further inquiries can be directed to the corresponding author.
Author contributions
KK: Investigation, Writing – original draft, Data curation. HK: Writing – review and editing, Supervision, Conceptualization.
Funding
The author(s) declared that financial support was not received for this work and/or its publication.
Acknowledgments
The authors are grateful to CIF, Lovely Professional University, Phagwara, for providing instrumental facilities that enable them to complete their research.
Conflict of interest
The author(s) declared that this work was conducted in the absence of any commercial or financial relationships that could be construed as a potential conflict of interest.
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Supplementary material
The Supplementary Material for this article can be found online at: https://www.frontiersin.org/articles/10.3389/fphy.2026.1874615/full#supplementary-material
References
1.
LitchfieldSGSchulzKGKelaherBP. The influence of plastic pollution and ocean change on detrital decomposition. Mar Pollut Bull (2020) 158:111354. 10.1016/j.marpolbul.2020.111354
2.
KaurKKaurH. Removal of microplastic contaminants from aquatic environment. In: Microplastic Pollution: Causes, Effects and Control. The Netherlands: Bentham Science Publishers (2023). p. 69–92. 10.2174/9789815165104123010007
3.
PramanikBKPramanikSKMoniraS. Understanding the fragmentation of microplastics into nano-plastics and removal of nano/microplastics from wastewater using membrane, air flotation and nano-ferrofluid processes. Chemosphere (2021) 282:131053. 10.1016/j.chemosphere.2021.131053
4.
BriassoulisD. Agricultural plastics as a potential threat to food security, health, and environment through soil pollution by microplastics: problem definition. Sci Total Environment (2023) 892:164533. 10.1016/j.scitotenv.2023.164533
5.
BhagwatKPRodrigueDRomero-ZerónL. Effective removal of microplastic particles from wastewater using hydrophobic bio-substrates. Pollutants (2024) 4(2):231–50. 10.3390/pollutants4020015
6.
JiangYZhangHHongLShaoJZhangBYuJet alAn integrated plasma–photocatalytic system for upcycling of polyolefin plastics. ChemSusChem (2023) 16(14):e202300106. 10.1002/cssc.202300106
7.
TeutenELRowlandSJGallowayTSThompsonRC. Potential for plastics to transport hydrophobic contaminants. Environ Science & Technology (2007) 41(22):7759–64. 10.1021/es071737s
8.
BesselingEQuikJTSunMKoelmansAA. Fate of nano-and microplastic in freshwater systems: a modeling study. Environ Pollution (2017) 220:540–8. 10.1016/j.envpol.2016.10.001
9.
ParkCMChuKHHerNJangMBaaloushaMHeoJet alOccurrence and removal of engineered nanoparticles in drinking water treatment and wastewater treatment processes. Separat & Purif Rev (2017) 46(3):255–72. 10.1080/15422119.2016.1260588
10.
ZhangQXuEGLiJChenQMaLZengEYet alA review of microplastics in table salt, drinking water, and air: direct human exposure. Environ Science & Technology (2020) 54(7):3740–51. 10.1021/acs.est.9b04535
11.
NovotnaKCermakovaLPivokonskaLCajthamLTPivokonskyM. Microplastics in drinking water treatment–current knowledge and research needs. Sci Total Environment (2019) 667:730–40. 10.1016/j.scitotenv.2019.02.431
12.
JiGXingYYouT. Biochar as adsorbents for environmental microplastics and nanoplastics removal. J Environ Chem Eng (2024) 12:113377. 10.1016/j.jece.2024.113377
13.
XiaoYTianYXuWZhuJ. Photodegradation of microplastics through nanomaterials: insights into photocatalysts modification and detailed mechanisms. Materials (2024) 17(11):2755. 10.3390/ma17112755
14.
ShiYLiuPWuXShiHHuangHWangHet alInsight into chain scission and release profiles from photodegradation of polycarbonate microplastics. Water Res (2021) 195:116980. 10.1016/j.watres.2021.116980
15.
WuYYiRWangYZhangCZhengJNingPet alLight-driven degradation of microplastics: mechanisms, technologies, and future directions. J Hazard Mater Adv (2025) 17:100628. 10.1016/j.hazadv.2025.100628
16.
AsiltürkMŞenerŞ. TiO2-activated carbon photocatalysts: preparation, characterization and photocatalytic activities. Chem Eng J (2012) 180:354–63. 10.1016/j.cej.2011.11.045
17.
HelaliSPuzenatEPerolNSafiMJGuillardC. Methylamine and dimethylamine photocatalytic degradation-adsorption isotherms and kinetics. Appl Catal A: Gen (2011) 402(1-2):201–7. 10.1016/j.apcata.2011.06.004
18.
VegaAAImoberdorfGEMohseniM. Photocatalytic degradation of 2, 4-dichlorophenoxyacetic acid in a fluidized bed photoreactor with composite template-free TiO2 photocatalyst. Appl Catal A: Gen (2011) 405(1-2):120–8. 10.1016/j.apcata.2011.07.033
19.
RafqahSWong-Wah-ChungPNelieuSEinhornJSarakhaM. Phototransformation of triclosan in the presence of TiO2 in aqueous suspension: mechanistic approach. Appl Catal B: Environ (2006) 66(1-2):119–25. 10.1016/j.apcatb.2006.03.004
20.
Fernández-IbáñezPIBlancoJMalatoSDe Las NievesFJ. Application of the colloidal stability of TiO2 particles for recovery and reuse in solar photocatalysis. Water Res (2003) 37(13):3180–8. 10.1016/S0043-1354(03)00157-X
21.
LargitteLPasquierRJCER. A review of the kinetics adsorption models and their application to the adsorption of lead by an activated carbon. Chem Engineering Research Design (2016) 109:495–504. 10.1016/j.cherd.2016.02.006
22.
OkonjiSOYuLDominicJAPernitskyDAchariG. Adsorption by granular activated carbon and nano zerovalent iron from wastewater: a study on removal of selenomethionine and selenocysteine. Water (2020) 13(1):23. 10.3390/w13010023
23.
KaurKKaurRKaurH. A systematic review of lignocellulosic biomass for remediation of environmental pollutants. Appl Surf Sci Adv (2024) 19:100547. 10.1016/j.apsadv.2023.100547
24.
LimTTYapPSSrinivasanMFaneAG. TiO2/AC composites for synergistic adsorption-photocatalysis processes: present challenges and further developments for water treatment and reclamation. Crit Rev Environ Sci Technol (2011) 41(13):1173–230. 10.1080/10643380903488664
25.
JjagweJOlupotPWMenyaEKalibbalaHM. Synthesis and application of granular activated carbon from biomass waste materials for water treatment: a review. J Bioresources Bioproducts (2021) 6(4):292–322. 10.1016/j.jobab.2021.03.003
26.
LiuZSunYXuXMengXQuJWangZet alPreparation, characterization and application of activated carbon from corn cob by KOH activation for removal of Hg (II) from aqueous solution. Bioresour Technol (2020) 306:123154. 10.1016/j.biortech.2020.123154
27.
ObeyGAdelaideMRamarajR. Biochar derived from non-customized matamba fruit shell as an adsorbent for wastewater treatment. J Bioresources Bioproducts (2022) 7(2):109–15. 10.1016/j.jobab.2021.12.001
28.
MaHXuZWangWGaoXMaH. Adsorption and regeneration of leaf-based biochar for p-nitrophenol adsorption from aqueous solution. RSC Advances (2019) 9(67):39282–93. 10.1039/C9RA07943B
29.
MinaeiSBenisKZMcPhedranKNSoltanJ. Evaluation of a ZnCl2-modified biochar derived from activated sludge biomass for adsorption of sulfamethoxazole. Chem Eng Res Des (2023) 190:407–20. 10.1016/j.cherd.2022.12.038
30.
AroraMKaurH. Effect of doping in TiO2/chitosan composite on adsorptive-photocatalytic removal of gallic acid from water. Chemosphere (2025) 373:144122. 10.1016/j.chemosphere.2025.144122
31.
AksoyEDemirNVarlikliC. White LED light production using dibromoperylene derivatives in Down conversion of energy. Can J Phys (2018) 96(7):734–9. 10.1139/cjp-2017-0752
32.
KaurKKaurH. Visible light induced photocatalytic degradation of polyvinyl chloride from water using graphene oxide/nitrogen-doped TiO2 nanocomposite. J Environ Chem Eng (2025) 13:115664. 10.1016/j.jece.2025.115664
33.
BhargavaSChuJJHValiyaveettilS. Controlled dye aggregation in sodium dodecylsulfate-stabilized poly (methylmethacrylate) nanoparticles as fluorescent imaging probes. ACS Omega (2018) 3(7):7663–72. 10.1021/acsomega.8b00785
34.
ThongkruaSSuriyaP. Removal of colour and COD in biologically pre-treated leachate using activated carbon from corn cobs. Pollution (2022) 8(2):657–70. 10.22059/poll.2022.328345.1153
35.
YanLLiuYZhangYLiuSWangCChenWet alZnCl2 modified biochar derived from aerobic granular sludge for developed microporosity and enhanced adsorption to tetracycline. Bioresour Technol (2020) 297:122381. 10.1016/j.biortech.2019.122381
36.
AretiHAJabesaAMuletaMDEmanaAN. Adsorptive performances and valorization of green synthesized biochar–based activated carbon from banana peel and corn cob composites for the abatement of Cr (VI) from synthetic solutions: parameters, isotherms, and remediation studies. Heliyon (2024) 10(13):e33811. 10.1016/j.heliyon.2024.e33811
37.
ZhangHChenBBanfieldJF. The size dependence of the surface free energy of titania nanocrystals. Phys Chem Chem Phys (2009) 11(14):2553–8. 10.1039/B819623K
38.
HeXChenXWangXJiangL. Optimization of activated carbon production from corn cob using response surface methodology. Front Environ Sci (2023) 11:1105408. 10.3389/fenvs.2023.1105408
39.
ParvathirajaCKatheriaSSiddiquiMRWabaidurSMIslamMALaiWC. Activated carbon-loaded titanium dioxide nanoparticles and their photocatalytic and antibacterial investigations. Catalysts (2022) 12(8):834. 10.3390/catal12080834
40.
LuLShanRShiYWangSYuanH. A novel TiO2/biochar composite catalysts for photocatalytic degradation of methyl orange. Chemosphere (2019) 222:391–8. 10.1016/j.chemosphere.2019.01.132
41.
XieMJingLZhouJLinJFuH. Synthesis of nanocrystalline anatase TiO2 by one-pot two-phase separated hydrolysis-solvothermal processes and its high activity for photocatalytic degradation of rhodamine B. J Hazardous Materials (2010) 176(1-3):139–45. 10.1016/j.jhazmat.2009.11.008
42.
XuNShiZFanYDongJShiJHuMZC. Effects of particle size of TiO2 on photocatalytic degradation of methylene blue in aqueous suspensions. Ind & Eng Chem Res (1999) 38(2):373–9. 10.1021/ie980378u
43.
WangXSøLSuRWendtSHaldPMamakhelAet alThe influence of crystallite size and crystallinity of anatase nanoparticles on the photo-degradation of phenol. J Catalysis (2014) 310:100–8. 10.1016/j.jcat.2013.04.022
44.
DeviLGMurthyBNKumarSG. Photocatalytic activity of TiO2 doped with Zn2+ and V5+ transition metal ions: influence of crystallite size and dopant electronic configuration on photocatalytic activity. Mater Sci Eng B (2010) 166(1):1–6. 10.1016/j.mseb.2009.09.008
45.
WangXZhangKGuoXShenGXiangJ. Synthesis and characterization of N-doped TiO2 loaded onto activated carbon fiber with enhanced visible-light photocatalytic activity. New J Chem (2014) 38(12):6139–46. 10.1039/C4NJ00962B
46.
FazalTRazzaqAJavedFHafeezARashidNAmjadUSet alIntegrating adsorption and photocatalysis: a cost effective strategy for textile wastewater treatment using hybrid biochar-TiO2 composite. J Hazardous Materials (2020) 390:121623. 10.1016/j.jhazmat.2019.121623
47.
Megat Nabil MohsinSHusseinMZSarijoSHFakuraziSArulselvanPTaufiq-YapYH. Nanolayered composite with enhanced ultraviolet ray absorption properties from simultaneous intercalation of sunscreen molecules. Int J Nanomedicine (2018) 13:6359–74. 10.2147/IJN.S171390
48.
AbdullahMIqbalJRehmanMSUKhalidUMateenFArshadSNet alRemoval of ceftriaxone sodium antibiotic from pharmaceutical wastewater using an activated carbon based TiO2 composite: adsorption and photocatalytic degradation evaluation. Chemosphere (2023) 317:137834. 10.1016/j.chemosphere.2023.137834
49.
WangSColeISLiQ. Quantum-confined bandgap narrowing of TiO2 nanoparticles by graphene quantum dots for visible-light-driven applications. Chem Commun (2016) 52(59):9208–11. 10.1039/C6CC03302D
50.
WoanKPyrgiotakisGSigmundW. Photocatalytic carbon-nanotube–TiO2 composites. Adv Mater (2009) 21(21):2233–9. 10.1002/adma.200802738
51.
EskandarlooHBadieiABehnajadyMATavakoliAZiaraniGM. Ultrasonic-assisted synthesis of Ce doped cubic–hexagonal ZnTiO3 with highly efficient sonocatalytic activity. Ultrason Sonochem (2016) 29:258–69. 10.1016/j.ultsonch.2015.10.004
52.
HuangYQinJFanZWeiDSeoHJ. Photoenergy conversion behaviors of photoluminescence and photocatalysis in silver-coated LiBaPO4: Eu2+. Inorg Chem (2019) 58(19):13161–9. 10.1021/acs.inorgchem.9b02037
53.
WangYLaiQZhangFShenXFanMHeYet alHigh efficiency photocatalytic conversion of CO2 with H2O over Pt/TiO2 nanoparticles. RSC Adv (2014) 4(84):44442–51. 10.1039/C4RA07457B
54.
LiuGYangHGPanJYangYQLuGQChengHM. Titanium dioxide crystals with tailored facets. Chem Reviews (2014) 114(19):9559–612. 10.1021/cr400621z
55.
ZhangXCaoSWuZZhaoSPiaoL. Enhanced photocatalytic activity towards degradation and H2 evolution over one dimensional TiO2@MWCNTs heterojunction. Appl Surf Sci (2017) 402:360–8. 10.1016/j.apsusc.2017.01.096
56.
AoWQuJYuHLiuYLiuCFuJet alTiO2/activated carbon synthesized by microwave-assisted heating for tetracycline photodegradation. Environ Res (2022) 214:113837. 10.1016/j.envres.2022.113837
57.
DjellabiRYangBWangYCuiXZhaoX. Carbonaceous biomass-titania composites with Ti-O-C bonding bridge for efficient photocatalytic reduction of Cr (VI) under narrow visible light. Chem Eng J (2019) 366:172–80. 10.1016/j.cej.2019.02.035
58.
KimJRKanE. Heterogeneous photocatalytic degradation of sulfamethoxazole in water using a biochar-supported TiO2 photocatalyst. J Environ Manage (2016) 180:94–101. 10.1016/j.jenvman.2016.05.016
59.
KlüpfelLKeiluweitMKleberMSanderM. Redox properties of plant biomass-derived black carbon (biochar). Environ Science & Technology (2014) 48(10):5601–11. 10.1021/es500906d
60.
LehmannJCzimczikCLairdDSohiS. Stability of biochar in soil. In: Biochar for Environmental Management. London, United Kingdom: Routledge, Taylor and Francis (2012). p. 215–38. 10.4324/9781849770552
61.
YangHYanRChenHLeeDHZhengC. Characteristics of hemicellulose, cellulose and lignin pyrolysis. Fuel (2007) 86(12-13):1781–8. 10.1016/j.fuel.2006.12.013
62.
GuptaGKRamMBalaRKapurMMondalMK. Pyrolysis of chemically treated corncob for biochar production and its application in Cr (VI) removal. Environ Prog & Sustain Energ (2018) 37(5):1606–17. 10.1002/ep.12838
63.
MinzPMishraS. Jackfruit peel derived ZnCl2-impregnated activated carbon: optimization, characterization, and application in dye removal. Biomass Convers Biorefinery (2023) 15:1–18. 10.1007/s13399-023-04592-0
64.
SaranyaKSVellora Thekkae PadilVSenanCPilankattaRSaranyaKGeorgeBet alGreen synthesis of high temperature stable anatase titanium dioxide nanoparticles using gum kondagogu: characterization and solar driven photocatalytic degradation of organic dye. Nanomaterials (2018) 8(12):1002. 10.3390/nano8121002
65.
LeeXJLeeLYGanSThangalazhy-GopakumarSNgHK. Biochar potential evaluation of palm oil wastes through slow pyrolysis: thermochemical characterization and pyrolytic kinetic studies. Bioresour Technology (2017) 236:155–63. 10.1016/j.biortech.2017.03.105
66.
BoujibarOSouiknyAGhamoussFAchakODahbiMChafikT. CO2 capture using N-containing nanoporous activated carbon obtained from argan fruit shells. J Environ Chem Eng (2018) 6(2):1995–2002. 10.1016/j.jece.2018.03.005
67.
WangCLHwangWSChuHLLinHJKoHHWangMC. Kinetics of anatase transition to rutile TiO2 from titanium dioxide precursor powders synthesized by a sol-gel process. Ceramics Int (2016) 42(11):13136–43. 10.1016/j.ceramint.2016.05.101
68.
PangYLLimSLeeRKL. Enhancement of sonocatalytic degradation of organic dye by using titanium dioxide (TiO2)/activated carbon (AC) derived from oil palm empty fruit bunch. Environ Sci Pollut Res (2020) 27(28):34638–52. 10.1007/s11356-019-05373-x
69.
SmithMScudieroLEspinalJMcEwenJSGarcia-PerezM. Improving the deconvolution and interpretation of XPS spectra from chars by ab initio calculations. Carbon (2016) 110:155–71. 10.1016/j.carbon.2016.09.012
70.
YueYYueXTangXHanLWangJWangSet alSynergistic adsorption and photocatalysis study of TiO2 and activated carbon composite. Heliyon (2024) 10(10):e30817. 10.1016/j.heliyon.2024.e30817
71.
ChenYWangCChenJWangSJuJKangW. Preparing biomass carbon fiber derived from waste rabbit hair as a carrier of TiO2 for photocatalytic degradation of methylene blue. Polymers (2022) 14(8):1593. 10.3390/polym14081593
72.
Castilla-CaballeroDHernandez-RamirezAVazquez-RodriguezSColina-MárquezJMachuca-MartínezFBarraza-BurgosJet alEffect of pyrolysis, impregnation, and calcination conditions on the physicochemical properties of TiO2/Biochar composites intended for photocatalytic applications. J Environ Chem Eng (2023) 11(3):110274. 10.1016/j.jece.2023.110274
73.
MaoCCWengHS. Promoting effect of adding carbon black to TiO2 for aqueous photocatalytic degradation of methyl orange. Chem Eng J (2009) 155(3):744–9. 10.1016/j.cej.2009.09.016
74.
NguyenTPTranQBLyQVHaiLTLeDTTranMBet alEnhanced visible photocatalytic degradation of diclofen over N-doped TiO2 assisted with H2O2: a kinetic and pathway study. Arabian Journal Chemistry (2020) 13(11):8361–71. 10.1016/j.arabjc.2020.05.023
75.
AbduMWorkuABabaeeSDialePMsagatiTANureJF. The development of TiO2-biochar composite material for photodegradation of basic blue 41 and erichrome black T azo dyes from water. Surf Inter (2025) 62:106156. 10.1016/j.surfin.2025.106156
76.
JedynakKCharmasB. Adsorption properties of biochars obtained by KOH activation. Adsorption (2024) 30(2):167–83. 10.1007/s10450-023-00399-7
77.
ChakhtounaHBenzeidHZariNel kacem QaissABouhfidR. Functional CoFe2O4‐modified biochar derived from banana pseudostem as an efficient adsorbent for the removal of amoxicillin from water. Separat Purif Technol (2021) 266:118592. 10.1016/j.seppur.2021.118592
78.
VarelaCFMoreno-AldanaLCAgámez-PertuzYY. Adsorption of pharmaceutical pollutants on ZnCl2-activated biochar from corn cob: efficiency, selectivity and mechanism. J Bioresources Bioproducts (2024) 9(1):58–73. 10.1016/j.jobab.2023.10.003
79.
DuncanSEHannahS. Light-Protective Packaging Materials for Foods and Beverages. Emerging Food Packaging Technologies. United Kingdom: Woodhead publishing (2012). p. 303–22. 10.1533/9780857095664.3.303
80.
ThomsLJGirwidzRMunichLMU. Experimenting from a distance: optical spectrometry via the Internet. MPTL18–B. Proc (2013) 59-63.
81.
PaiuMLuticDFavierLGavrilescuM. Heterogeneous photocatalysis for advanced water treatment: materials, mechanisms, reactor configurations, and emerging applications. Appl Sci (2025) 15(10):5681. 10.3390/app15105681
82.
DorrajMAlizadehMSairiNABasirunWJGohBTWoiPMet alEnhanced visible light photocatalytic activity of copper-doped titanium oxide–zinc oxide heterojunction for methyl orange degradation. Appl Surf Sci (2017) 414:251–61. 10.1016/j.apsusc.2017.04.045
83.
ZhaoXXLiuDLiYHCuiGH. Two 3D cadmium (II) coordination polymers modulated by flexible bis (benzimidazole) ligands displaying high photocatalytic activities for degradation of methylene blue and methyl orange. Polyhedron (2018) 156:200–7. 10.1016/J.POLY.2018.09.041
84.
BakhshianSSahimiM. Adsorption-induced swelling of porous media. Int J Greenhouse Gas Control (2017) 57:1–13. 10.1016/j.ijggc.2016.12.011
85.
BatoolAValiyaveettilS. Coprecipitation—An efficient method for removal of polymer nanoparticles from water. ACS Sustain Chem & Eng (2020) 8(35):13481–7. 10.1021/acssuschemeng.0c04511
86.
DE LorenziAGiorgianniSBiniR. High-resolution FTIR spectroscopy of the C—Cl stretching mode of vinyl chloride. Mol Phys (1999) 96(1):101–8. 10.1080/00268979909482942
87.
MiaoFLiuYGaoMYuXXiaoPWangMet alDegradation of polyvinyl chloride microplastics via an electro-fenton-like system with a TiO2/graphite cathode. J Hazard Mater (2020) 399:123023. 10.1016/j.jhazmat.2020.123023
88.
BashirMAhangerMAGaniKM. Investigations on adsorptive removal of PVC microplastics from aqueous solutions using pinus roxburghii–derived biochar. Environ Sci Pollut Res (2024) 31(49):59416–29. 10.1007/s11356-024-35166-w
89.
TiltonRDRobertsonCRGastAP. Manipulation of hydrophobic interactions in protein adsorption. Langmuir (1991) 7(11):2710–8. 10.1021/la00059a051
90.
NosratabadNAYanQCaiZWanC. Exploring nanomaterial-modified biochar for environmental remediation applications. Heliyon (2024) 10:e37123. 10.1016/j.heliyon.2024.e37123
91.
WeidnerEKarbassiyazdiEAltaeeAJesionowskiTCiesielczykF. Hybrid metal oxide/biochar materials for wastewater treatment technology: a review. ACS Omega (2022) 7(31):27062–78. 10.1021/acsomega.2c02909
92.
ZhengHChenQChenZ. Carbon-based adsorbents for micro/nano-plastics removal: current advances and perspectives. Water Emerging Contaminants & Nanoplastics (2024) 3(2):N–A. 10.20517/wecn.2023.74
93.
Valiyaveettil SalimkumarAKurisingal CleetusMCEhigieJOOnogboseleCONishaPKumarBSet alAdsorption behavior and interaction of micro-nanoplastics in soils and aquatic environment. Management of micro and nano-plastics in soil and biosolids: fate, occurrence. Monit Remedies (2024) 283–311. 10.1007/978-3-031-51967-3_11
94.
DasJYadavEPoluriKM. A review on the role of nanotechnological interventions in sequestration, mitigation and value-added product conversion of micro-/nanoplastics. Environ Sci Nano (2025) 12:189–218. 10.1039/D4EN00267A
95.
MulindwaPKasuleJSNantabaFWasswaJExpósitoAJ. Bioadsorbents for removal of microplastics from water ecosystems: a review. Int J Sustain Eng (2024) 17(1):582–99. 10.1080/19397038.2024.2374003
96.
AlomNRoyTSarkarTRaselMHossainMSJamalM. Removal of microplastics from aqueous media using activated jute stick charcoal. Heliyon (2024) 10(18):e37380. 10.1016/j.heliyon.2024.e37380
97.
WuPCaiZJinHTangY. Adsorption mechanisms of five bisphenol analogues on PVC microplastics. Sci Total Environ (2019) 650:671–8. 10.1016/j.scitotenv.2018.09.049
98.
LiangSXuSWangCLingJXianZWuHet alEnhanced alteration of poly (vinyl chloride) microplastics by hydrated electrons derived from indole-3-acetic acid assisted by a common cationic surfactant. Water Res (2021) 191:116797. 10.1016/j.watres.2020.116797
99.
CuiJZhangFLiHCuiJRenYYuX. Recent progress in biochar-based photocatalysts for wastewater treatment: synthesis, mechanisms, and applications. Appl Sci (2020) 10(3):1019. 10.3390/app10031019
100.
YangLLuoYYangLLuoSLuoXDaiWet alEnhanced photocatalytic activity of hierarchical titanium dioxide microspheres with combining carbon nanotubes as “e-bridge”. J Hazardous Materials (2019) 367:550–8. 10.1016/j.jhazmat.2019.01.016
101.
ZarrinSHeshmatpourF. Photocatalytic activity of TiO2/Nb2O5/PANI and TiO2/Nb2O5/RGO as new nanocomposites for degradation of organic pollutants. J Hazardous Materials (2018) 351:147–59. 10.1016/j.jhazmat.2018.02.052
102.
LiuXYangYShiXLiK. Fast photocatalytic degradation of methylene blue dye using a low-power diode laser. J Hazardous Materials (2015) 283:267–75. 10.1016/j.jhazmat.2014.09.031
103.
HirakawaTKamatPV. Charge separation and catalytic activity of ag@ TiO2 core-shell composite clusters under UV− irradiation. J Am Chem Soc (2005) 127(11):3928–34. 10.1021/ja042925a
104.
XiongZWGuoLNJiangHZhangYGaoQLiJet alA handy way for forming N-doped TiO2/carbon from pectin and d, l-serine hydrazide hydrochloride. Int J Biol Macromolecules (2024) 277:134155. 10.1016/j.ijbiomac.2024.134155
105.
LiDQMengYJLiJSongYJXuF. TiO2/carbonaceous nanocomposite from titanium-alginate coordination compound. Carbohydr Polymers (2022) 288:119400. 10.1016/j.carbpol.2022.119400
Summary
Keywords
adsorption, corncob activated carbon, nitrogen doped titanium oxide, photocatalysis, polyvinyl chloride degradation
Citation
Kaur K and Kaur H (2026) Biomass-based N-TiO2 composites for polyvinylchloride nano plastic photodegradation. Front. Phys. 14:1874615. doi: 10.3389/fphy.2026.1874615
Received
07 May 2026
Revised
16 June 2026
Accepted
11 August 2026
Published
04 September 2026
Volume
14 - 2026
Edited by
A. K. Haghi, University of Coimbra, Portugal
Reviewed by
Halyna Bodnar Yankovych, Slovak Academy of Sciences (SAS), Slovakia
De-qiang Li, Xinjiang Agricultural University, China
Updates
Copyright
© 2026 Kaur and Kaur.
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*Correspondence: Harpreet Kaur, harpreet2.kaur@lpu.co.in
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