In this work, the formation of carbon-based nanomaterials–fulvic acid (CNMs-FA) composites and their capacities for the adsorption and photodegradation of typical organic contaminants in aqueous solutions were investigated. The results suggested that the formation of CNMs-FA composites was dominated by adsorbing FA on CNMs via the physisorption process, which fit the pseudo-first-order kinetic model and the Langmuir isotherm model. The formed CNMs-FA composites were characterized by using the Brunauer–Emmett–Teller, scanning electron microscopy, and infrared spectroscopy techniques and further applied for examining their effects on the adsorption and photodegradation of selected organic contaminants in aqueous solutions. The adsorption of organic contaminants on CNMs-FA composites is mainly involved in hydrogen bonding and electrostatic interactions between organic contaminants and FA species adhering to CNMs. In addition, the CNMs-FA composites are able to promote the photosensitive degradation of organic contaminants due to the photogenerated reactive species including ROS and CNMs-3FA* under sunlight irradiation. This study provided a deeper and more comprehensive understanding of the environmental behavior of CNMs in real natural surface water and clarified the underlying mechanisms.

  • Carbon-based nanomaterials (CNMs) preferentially adsorb fulvic acid to form CNMs-FA in environmental water.

  • Adsorption of FA on CNMs fits pseudo-first-order kinetic and Langmuir isotherm models.

  • CNMs-FA can adsorb organic pollutants via hydrogen bonding and electrostatic interactions.

  • CNMs-FA promotes the photodegradation of organic pollutants under sunlight irradiation.

Carbon-based nanomaterials (CNMs), such as graphene oxide (GO), single-wall carbon nanotube (SWCNT), and multi-wall carbon nanotube (MWCNT), have been widely applied in various industrial and agricultural products and even daily supplies due to their unique chemical and physical properties, including high surface area and adsorption capacity, superb electrical and thermal conductivity, high mechanical strength, cost-effectiveness of mass production, nanoscale size effects, etc. (Zhang et al. 2022; Pérez et al. 2023). For example, both CNTs and magnetic GO have been utilized as active components of adsorbents for removing the various inorganic and organic pollutants from industrial wastewater and landfill leachate (Lerman et al. 2013; Zhang et al. 2016; Li et al. 2020). Meanwhile, during the production, transportation, utilization, and disposal of the CNMs or CNMs-contained products, CNMs are inevitably released into the aquatic environment through domestic sewage, industrial wastewater, and even coastal recreation activities (Park et al. 2017; Sharma et al. 2020; Ding et al. 2022). Therefore, the ecological risk and environmental behavior of CNMs in the aquatic ecosystem have received increasing concern from the scientific community (Qiang et al. 2016; De Marchi et al. 2018; Ko et al. 2019). Previous studies on the biological impacts of CNMs on aquatic ecology have shown that CNMs could inhibit algal growth, cause mechanical damage and oxidative stress, and induce light shielding effects, but promote the production of bioactive substances and prevention of biological contaminants (Jackson et al. 2013; Wang et al. 2019; Huang et al. 2022). Other studies suggested that CNMs may cause incubation delay, cardiac edema, chorionic processes and changes in the protein structure of aquatic animals, as well as remarkably enhance the accumulation of organic contaminants such as perfluoro-octane sulfanilamide (PFOS) in their blood, guts, and muscles (Chen et al. 2015; Qiang et al. 2016). As for the environmental behavior of CNMs in aquatic environments, most previous investigations focused on their aggregation and adsorption of environmental contaminants including heavy metals and organic contaminants (Apul et al. 2013; Jiang et al. 2017a; Ighalo et al. 2022; Rout et al. 2023). However, due to the abundance of dissolved organic matter (DOM) in environmental water, it is necessary to take into account the naturally occurring DOM for comprehensively understanding the environmental behavior of CNMs in a real aquatic environment (Jiang et al. 2017a, 2017b). Many studies demonstrated that the CNMs can easily adsorb DOM by hydrophobic interaction, electrostatic interaction, hydrogen bonding, and π–π interaction (Goodwin et al. 2018; Ren et al. 2018; Zhou et al. 2019). Furthermore, a few reports implied that the CNMs prefer to adsorb the DOM to form CNMs-DOM composites, leading to a significant reduction of the adsorption capacity to coexisting contaminants by blocking the adsorption sites on CNMs’ surface, probably due to the much higher concentrations of DOM than those of trace pollutants in aquatic environments (Lerman et al. 2013; Zhang et al. 2016; Munoz et al. 2021). On the other hand, it is well known that the DOM is the most photochemically active fraction of species mainly contributing to the indirect photodegradation of persistent organic pollutants (POPs) in environmental surface water since they can predominantly absorb sunlight irradiation and produce abundant reactive species such as 1O2, OH, and triplet 3DOM* (Yu et al. 2019). However, it is unknown what the impacts of CNMs-DOM composites on the photodegradation of organic contaminants in environmental surface water until now. Thus, investigating the formation of CNMs-DOM composites and their capacities for the adsorption and photodegradation of organic contaminants would be vital to clarifying the effects of CNMs on the migration and transformation of environmental pollutants in real natural surface water.

Herein, the aims of this study were to explore the adsorption and photochemical decomposition of organic contaminants on CNMs in the presence of DOM at naturally occurring concentrations under sunlight irradiation. Fulvic acid (FA) was selected as a surrogate of natural DOM since FA is the most important and abundant component of natural DOM in aquatic environments (Yu et al. 2019; Jiao et al. 2022). Three typical CNMs (SWCNT, MWCNT, and GO) and three typical emerging organic contaminants, including estradiol, dibutyl phthalate (DBP), and tetracycline (TC) (chemical structures are presented in Supplementary Figure S1), were selected to conduct the adsorption and photodegradation experiments. The adsorption kinetics and isotherms of FA on CNMs were determined and the factors affecting the adsorption processes, such as pH and concentration, were evaluated. The prepared CNMs-FA composites were characterized and adsorption and photodegradation processes of the interested organic contaminants on the surface of CNMs-FA composites were investigated.

Chemicals and materials

Multi-walled carbon nanotubes (MWCNTs, >99%), single-walled carbon nanotubes (SWCNTs, >95%), DBP, and furfuryl alcohol (FFA) were obtained from Tokyo Chemical Industry Co., Ltd (Tokyo, Japan). Graphene oxide (GO, >98%), ethinyl estradiol (EE2), 2,4,6-trimethylphenol (TMP, 98%), isopropyl alcohol (IPA, 99.8%), and fulvic acid (FA, 90.0%) were supplied by Aladdin Industrial Co., Ltd (Shanghai, China). TC, sodium hydroxide (NaOH), and phosphoric acid (H3PO4) were purchased from Sinopharm Chemical Reagent Co., Ltd (Shanghai, China). HPLC-grade methanol (MeOH) and acetonitrile (ACN) were supplied by Supelco (USA). Except where noted, all reagents were of analytical grade and distill-deionized water (18.3 mΩ cm) was applied in all experiments.

Adsorption of FA with CNMs

Before studying the adsorption and photodegradation of organic contaminants with CNMs-FA composites, the adsorption performance of FA with CNMs, including adsorption capacities, kinetics, and isotherms, was comprehensively studied by static adsorption experiments. The investigations were performed in 100 mL conical flasks containing 100 mL FA solution and 5.0 mg CNMs (dry weight) and the mixture was stirred at 120 rpm in a water bath at 25 ± 1 °C under dark conditions. The pH of the aqueous FA solution was adjusted to 5.0, 7.0, and 9.0 using 0.1M HClO4 and 0.1M NaOH solutions. Adsorption kinetic experiments were performed in conical flasks containing aqueous FA solutions (initial concentration: 16.0 mg L−1) and pre-dried CNMs (5.0 mg) at pH 7.0 ± 0.1. Additionally, isothermal experiments were conducted at the initial concentrations of the FA solution in the range of 2.0–16.0 mg L−1 at pH 7.0 ± 0.1 and stirred at 120 rpm for 12 h at 25 ± 1 °C. The concentration of FA in sample solutions was measured by using UV–Vis spectroscopy (UV-1900, Shimadzu, Japan) at 285 nm via the calibration curve method. After the adsorption process, the suspensions were centrifuged at 1 × 104 rpm for 15 min and filtered with 0.22 μm polytetrafluoroethylene (PTFE) membrane. Then, the filtrated CNMs-FA composites were washed with distill-deionized water until the residual FA in the eluent could not be detected. Finally, the CNMs-FA composites were dried in a vacuum freeze dryer for 24 h and stored in a desiccator for further utilization.

Characterization of CNMs and CNMs-FA

The original CNMs and prepared CNMs-FA composites were characterized by several techniques to confirm the formation of CNMs-FA and examine their physicochemical properties. A cold-field scanning electron microscope (SEM; SU8010, Hitachi, Japan) was used for morphological observation by uniformly attaching the samples to an aluminum block conductive glue. The infrared spectra were recorded by using a Fourier transform infrared spectroscopy (FTIR; Nicolet iS10, Thermo, USA). The porosity of the samples was determined at liquid nitrogen temperature using a surface area and porosity analyzer (Autosorb-iQ, Quantachrome, USA) after the evacuation of the samples at 150 °C for 12 h.

Adsorption of organic contaminants on CNMs-FA

To study the adsorption of organic contaminants by CNMs-FA composites, 30 mL of EE2, DBP, or TC at a concentration of 5.0 mg L−1 was added into a 50-mL quartz tube. Then, suspensions of 0.6 mL of 1.0 g L−1 CNMs-FA (SWCNT-FA, MWCNT-FA, or GO-FA) were added and stirred at 120 rpm for 12 h in a water bath at 25 ± 1 °C under dark. Then, they were centrifuged at 1 × 104 rpm for 15 min and the concentration of EE2, DBP, or TC in the supernatant was measured by using high-performance liquid chromatography (HPLC) by the calibration curve method. The adsorption efficiencies of studied contaminants on CNMs-FA were calculated by the concentration change of contaminants in samples before and after adsorption.

Photodegradation of organic contaminants in the presence of CNMs-FA

To examine the photodegradation of organic contaminants in the presence of CNMs-FA composites under solar irradiation, a merry-go-round photochemical chamber reactor equipped with a 500 W Xenon Lamp (XE-JY250, Jiuping Instrument Co., Ltd, Wuxi, China) was used for carrying out all photolysis experiments. The emission spectrum of the Xenon lamp solar simulator is shown in Supplementary Figure S2. The light below 290 nm from the Xenon lamp was filtered with a cutoff filter to simulate sunlight. The intensity of the lamp was set at 5.0 mW cm−2 to simulate the average intensity of the sunlight. Eight quartz photolysis tubes containing 20 mL of solution were held in the ring of the merry-go-round accessory. A hollow cylindrical lampshade with circulated cooling water (DC-1006, Wuxi Jiuping Co. Ltd, China) was employed to maintain the temperature at 25 ± 1 °C. The ring rotated within the reactor chamber at a speed of 20 rpm to give uniform irradiation to the quartz tubes. Tubes for the control samples were wrapped in aluminum foil to prevent light irradiation. The effect of pH on the photodecomposition of organic contaminants induced by CNMs-FA composites was evaluated. 5.0 mg L−1 of individual organic contaminant (EE2, DBP, and TC) and 200 mg L−1 of CNMs-FA (SWCNT-FA, MWCNT-FA, and GO-FA) suspensions were orthogonally prepared in 30 mL of aqueous solutions at a controlled pH of 7.0 ± 0.1 in quartz tubes. Then, the prepared sample solutions containing CNMs-FA and studied contaminants conducted photolysis experiments under simulated sunlight. Aliquots of samples were withdrawn after irradiating for 0.5, 1, 2, 4, and 8 h and centrifuged for 15 min at 1 × 104 rpm. Following filtration with a 0.22 μm filter, the samples were injected into HPLC to analyze the concentration of the studied contaminants.

Analytical methods

The concentrations of selected organic contaminants, including EE2, DBP, and TC, were analyzed by an HPLC (LC-20AT, Shimadzu, Tokyo, Japan) coupled with a UV–Vis detector (SPD-20A, Shimadzu, Tokyo, Japan) and a symmetry C18 column (150 × 3.9 mm, 5 μm particle size, Waters, USA) guarded by a 10 mm C18 guard column. The column temperature was maintained at 30 °C. The different mobile phases, flow rates, and detection wavelengths were applied for the analysis of different organic contaminants. For EE2, a mixture of ACN and deionized water (80:20; v/v) at a flow rate of 0.5 mL min−1 was employed as a mobile phase and the detection wavelength was set at 230 nm; for DBP, a mixture of MeOH and deionized water (90:10; v/v) at a flow rate of 1.0 mL min−1 was employed as a mobile phase and the detection wavelength was set at 275 nm; for TC, a mixture of methanol and 0.1% H3PO4 (35:65; v/v) at a flow rate of 1.0 mL min−1 was employed as a mobile phase and the detection wavelength was set at 350 nm. The identification and quantification of organic contaminants in samples were performed by matching retention time against that of the standard and the calibration equation method, respectively. The working standard solutions for three organic contaminants were prepared freshly in the concentration range of 0.00–50.0 mg L−1 by diluting the stock standards (1.0 mg mL−1), which were prepared by dissolving 20 mg of the corresponding standards in 20.0 mL of methanol.

Adsorption of FA on CNMs to form CNMs-FA composites

Effects of pH

The solution pH is a very important factor in the adsorption process because pH can affect the surface characteristics of the adsorbent and distribute adsorbate (Zhou et al. 2020). In this work, the effects of pH on the adsorption capacity of FA with CNMs were investigated in the pH range from 5.0 to 9.0, which are typical pH values of environmental surface water (Lingamdinne et al. 2022). The adsorption capacity was calculated according to the following equation:
formula
(1)
where C0 and Ce are the initial and equilibrium concentration of FA (mg L−1), respectively; qe is the amount of FA adsorbed at equilibrium (mg g−1); m is the mass of the CNMs (g); and V is the volume of the sample solution (L).

The uptake of FA by CNMs varied with solution pH, as shown in Supplementary Figure S3, and the absorption capacity of FA decreased from 84.93 to 30.05 mg g−1, from 48.65 to 5.51 mg g−1, and from 104.9 to 19.2 mg g−1 for SWCNT, MWCNT, and GO, respectively, when pH increased from 5.0 to 9.0 adjusted by 0.1M of HClO4 and NaOH solutions. This may be attributed to the enhanced hydrolysis of FA with the increase of pH, leading to the dehydrogenation of –OH and –COOH functional groups in FA, which resulted in the decrease of hydrogen-bonding donors, the abatement of hydrophobicity, and the increase of solubility in water (Huang et al. 2000; Lee et al. 2015). In addition, the negative charge associated with the ionized –OH and –COOH functional groups of FA may enhance the electrostatic repulsion between CNMs and FA. This can be confirmed by the observations that the surfaces of all three CNMs are negatively charged based on the results of Zeta potential measurements (Supplementary Figure S3), which were conducted by using a Zeta potential meter (Zetasizer Nano-ZS90, Malvern, UK) at different pHs from 5.0 to 9.0.

Adsorption kinetics and isotherms

Adsorption kinetic and isotherm analyses are reasonable tools to examine the adsorptive interaction between adsorbent and adsorbate (Liu et al. 2020). In this study, both kinetic and isotherm experiments for the adsorption of FA with CNMs were conducted by adding 0.05 g CNMs (dry weight) into 50 mL of aqueous FA solutions at pH 7.0 ± 0.1 and 25 ± 1 °C. Of those, kinetic experiments were carried out in time intervals of 0.5, 1, 2, 4, 8, and 12 h and isotherm experiments were performed with different initial concentrations of FA followed by stirring at 120 rpm for 12 h. The pseudo-first-order and pseudo-second-order models were applied to evaluate the adsorption kinetics of FA with CNMs according to the following Equations (2) and (3), respectively (Yang et al. 2014):
formula
(2)
formula
(3)
where qt is the adsorption capacity of FA at time t; qe,1 and qe,2 represent the adsorption capacity simulated by the pseudo-first-order kinetic model and the pseudo-second-order kinetic model, respectively; k1 and k2 are the pseudo-first-order kinetic adsorption rate (min−1) and the pseudo-second-order kinetic adsorption rate (g mg−1 min−1), respectively.
The adsorption isotherms of FA on CNMs were examined by using the Langmuir and Freundlich isotherm models based on Equations (4) and (5), respectively (Foo & Hameed 2010; Ighalo et al. 2022):
formula
(4)
formula
(5)
where qe and Ce are the adsorption capacity (mg g−1) and concentration of FA (mg L−1) at equilibrium time, respectively; qm is the maximum adsorption capacity of the CNMs (mg g−1); KL notes the Langmuir adsorption coefficient (L mg−1); Kf is the Freundlich affinity coefficient [(mg g−1)/(mg L−1)N]; and N is the empirical parameter.

As depicted in Supplementary Figure S4(a), the adsorption kinetic curves showing that the uptake of FA quickly increased in the initial 2 h and subsequently slowly changed until the adsorption process reached equilibrium. In addition, the data (Supplementary Table S1) verified that the pseudo-first-order model is more reasonable for kinetic data than the pseudo-second-order model for the adsorption of FA with all three CNMs according to the correlation coefficient (R2) values. As shown in Supplementary Figure S4(b) and presented in Supplementary Table S2, compared with the Freundlich model, the experimental results were more consistent with the Langmuir model according to the correlation coefficients (R2), suggesting that the adsorption process was monolayer adsorption. Furthermore, based on the calculated results via the Langmuir model, the maximum adsorption capacities (qm) of FA on SWCNT, MWCNT, and GO were 53.86 ± 6.85, 22.52 ± 2.31, and 34.65 ± 5.72 mg g−1, respectively, which are consistent with the data of specific surface areas and pore volumes analysis based on N2 adsorption and desorption for CNMs (Supplementary Table S3). Taken together, similar kinetic and isotherm models for the adsorption of FA on MWCNT, SWCNT, and GO were obtained, indicating that the similar adsorption mechanisms of the adsorption of FA with different CNMs were complied. The results revealed that the adsorption process of FA with CNMs was dominated by physisorption, which mainly involved hydrophobic interaction, Van der Waals force, π–π, and hydrogen-bonding interactions (Yang & Xing 2009; Jiang et al. 2017a; Zhou et al. 2019).

Characterization of CNMs-FA composites

SEM. The morphological images of the CNMs before and after the adsorption of FA were recorded by SEM, as presented in Figure 1(a)–1(f). Obviously, the pipe orifice margins of SWCNT and MWCNT were characterized by a smooth surface, with no covering on the surface and fewer agglomerations than SWCNT-FA and MWCNT-FA, respectively, whose images showed a layer of substance covering the surface of CNTs. The original GO has a smooth surface and translucence, curling gently and possessing slight folds. By contrast, both identical spots and agglomerations can be observed on the surface of GO-FA. These observations demonstrated that the FA has been adsorbed on the CNMs and further changed their surface morphology.
Figure 1

SEM images of (a) SWCNT, (b) MWCNT, (c) GO, (d) SWCNT-FA, (e) MWCNT-FA, (f) GO-FA, (g) nitrogen adsorption and desorption isotherms of CNMs and CNMs-FA, and (h and i) FTIR spectra of FA, CNMs, and CNMs-FA.

Figure 1

SEM images of (a) SWCNT, (b) MWCNT, (c) GO, (d) SWCNT-FA, (e) MWCNT-FA, (f) GO-FA, (g) nitrogen adsorption and desorption isotherms of CNMs and CNMs-FA, and (h and i) FTIR spectra of FA, CNMs, and CNMs-FA.

Close modal

BET N2 adsorption–desorption isotherm. The specific surface areas (SSAs) and pore volumes (PVs) of the studied CNMs and the corresponding CNMs-FA composites were measured by the N2 adsorption data collected for each sample at liquid nitrogen temperature (Bel Japan, Inc.) and further calculated using the standard Brunauer–Emmett–Teller (BET) method, respectively. As presented in Figure 1(g) and Supplementary Table S3, both SSABET and PV of CNMs-FA were significantly reduced by 9.8–23.8% and 7.0–23.7% compared to the original corresponding CNMs, respectively, indicating that the adsorption sites on the surface of CNMs were occupied and the porous structure of CNMs might be clogged due to the adsorption of FA.

FTIR. The FTIR is one of the most powerful techniques to characterize the presence of functional groups and chemical bonds in organic compounds and synthesized materials. In this study, the FTIR spectra of SWCNT, SWCNT-FA, MWCNT, MWCNT-FA, GO, GO-FA, and FA were recorded as shown in Figure 1(h) and 1(i). The SWCNT and MWCNT exhibit typical C = C stretching vibration at ∼1,600 cm−1. Meanwhile, the C = C peaks of SWCNT-FA and MWCNT-FA show red shifts of 1612 → 1639 and 1608 → 1627, respectively, which implies CNTs probably adsorbed FA via π–π interaction (Ahmed et al. 2018; Tokgöz et al. 2020). For GO, the bands at 3,407, 1,726, 1,396, 1,225, and 1,051 cm−1 are attributed to O–H, =C–O, carboxyl C–OH, epoxy CO–OC, and alkoxy CO–O stretching vibrations, respectively, corresponding to the sp3 structure of GO (Hartono et al. 2009). The band at 1,620 cm−1 is assigned to the C = C stretching vibration, corresponding to the sp2 structure of GO (Chen et al. 2021). The intensity of the C = C peak significantly increases in GO-FA than that in the original GO, indicating that the sp2 structure was enhanced due to the FA coating on GO. Also, the C = C peak of GO-FA shows a red shift of about 50 cm−1 compared to that of FA. The reason might be that the dual-band characteristics of the aromatic ring of FA can be weakened by π–π interaction (Chen et al. 2021). Additionally, O–H flexural vibration peak shifts from 3,407 to 3,379 cm−1 and C–O flexural vibration moves from 1,053 to 1,072 cm−1 after absorbing FA, which suggests hydrogen bonding between FA and GO (Jiang et al. 2017b; Zhou et al. 2019). These observations strongly confirmed that the FA was adsorbed on CNMs and the dominant interactions are π–π interaction and hydrogen bonding.

Adsorption of organic contaminants on CNMs-FA composites

Before the investigation of adsorption and photodegradation of organic contaminants on CNMs-FA composites, the HPLC analytical methods for selected organic contaminants, including EE2, DBP, and TC, were established and validated. The HPLC chromatograms and calibration equations for the identification and quantification of interested contaminants in samples are illustrated in Supplementary Figures S5 and S6, respectively. As shown in Figure 2, the adsorption processes of all three studied contaminants with three types of CNMs-FA quickly reached the equilibrium in 1 h under pH 7.0 ± 0.1 and 25 ± 1 °C. The adsorption capacities of all CNMs-FA composites for three pollutants followed the same order of DBP > EE2 > TC, indicating that the adsorption mechanisms of organic contaminants on different CNMs-FA are similar. In addition, the adsorption capacity of each of the pollutants with different types of CNMs-FA followed the order of SWCNT-FA > MWCNT-FA > GO-FA, which was consistent with the SSABET and PV data of these CNMs-FA composites (Supplementary Table S3). On the other hand, as presented in Supplementary Figure S7, the adsorption capacities on CNMs-FA for all three organic contaminants were much less than those on the corresponding CNMs. This may be attributed to that the microstructures and decreased SSABET of the CNMs-FA composites, as illustrated in Figure 1 and Supplementary Table S3, were expected to decrease their adsorption capabilities to the organic contaminants than the original CNMs since the adsorbents with uniform and more open structure are favorable for the adsorption of the coexistent substances (Awad et al. 2020; Khan et al. 2021).
Figure 2

Adsorption of organic contaminants with CNMs-FA composites at pH 7.0 ± 0.1 and 25 ± 1 °C. (a) EE2, (b) DBP, and (c) TC.

Figure 2

Adsorption of organic contaminants with CNMs-FA composites at pH 7.0 ± 0.1 and 25 ± 1 °C. (a) EE2, (b) DBP, and (c) TC.

Close modal
Furthermore, the results shown in Figure 3 implied that the pH of the solution and initial concentration of organic contaminants exhibited remarkable influences on the adsorption of organic contaminants on CNMs-FA composites. For the effects of initial concentration of organic contaminants, the results (Figure 3(d)–3(f)), that adsorption efficiencies of all three contaminants on three CNMs-FA decreased with the increase of initial concentration, can be well explained by the limited adsorption capacity of the given amount of CNMs-FA composites due to the limitation of available adsorption sites (Nguyen et al. 2021). When it comes to the effects of pH on adsorption processes between studied pollutants and CNMs-FA composites, more intricate results were obtained due to the more complex surface characteristics of the CNMs-FA composites than original CNMs and distribute organic contaminants containing a variety of hydrolyzable functional groups (Supplementary Figure S1). As for CNMs-FA composites, most of the surface of CNMs is mainly occupied by FA molecules as noted above. Thus, there is hydrogen bonding between CNMs-FA and organic contaminants since both FA and organic contaminants consist of diverse functional groups with hydrogen-bonding donors (Wu & Chen 2019). However, the –OH and –COOH functional groups of FA coating on the CNMs surface can be dehydrogenated into –O and –COO associated with a negative charge when the pH is over 4.0 (Huang et al. 2000). Furthermore, the negative charge may increase with the increase of pH on the surface of CNMs-FA, which was verified by the results of Zeta potential measurement (Supplementary Figure S8), due to the production of more –O and –COO anions. The molecule of EE2 can be protonated in the range of pH 5.0–9.0 since its pKa value is 10.73 (Sun & Zhou 2014). Thus, the adsorption of EE2 on CNMs-FA with negative charge gradually enhanced due to the electrostatic attraction, as shown in Figure 3(a). Meanwhile, the adsorption efficiency of DBP on CNMs-FA composites decreased with the increase of the solution pH (Figure 3(b)). This is likely to be caused by the fact that DBP would undergo quick hydrolysis in alkaline solutions to be transformed into monobutyl phthalate (MBP) or phthalic acid (Gao et al. 2016). As for TC, it is a kind of amphiprotic compound containing both acidic and alkaline functional groups with the pKa1, pKa2, and pKa3 of 3.3, 7.7, and 9.7, respectively (Zhao et al. 2021). Therefore, the hydrogenated form of TCH3+ is the prominent species resulting from the protonation of the dimethylamino functional group of TC in the pH < 3.3 and the amphiprotic forms of TCH2 and TC–H are dominant species in the pH range of 3.3–7.7. Meanwhile, TC mainly exists as alkaline forms of TC–H and TC2− when the solution pH increases over 7.7. Thus, there is an increasing electrostatic repulsion force between TC and CNMs-FA with the rise of pH, leading to the decreasing adsorption of TC on CNMs-FA composites (Figure 3(c)). In summary, the uptake of organic contaminants on CNMs-FA composites depends on the physicochemical properties of the pollutant itself and the pH of the aqueous solution. The adsorption process of studied contaminants on different CNMs-FA is similar and mainly dominated by hydrogen bonding and electrostatic interactions between organic contaminants and FA species adhering to CNMs.
Figure 3

Effects of pH and initial concentrations on the adsorption of typical organic contaminants with CNMs-FA composites. (a, d) EE2, (b, e) DBP, and (c, f) TC.

Figure 3

Effects of pH and initial concentrations on the adsorption of typical organic contaminants with CNMs-FA composites. (a, d) EE2, (b, e) DBP, and (c, f) TC.

Close modal

CNMs-FA mediated photodegradation of organic contaminants

It is well known that photodegradation, including direct photodegradation and indirect photosensitive degradation, is one of the most important naturally occurring transformation pathways of contaminants and plays a remarkable role in removing organic contaminants in environmental surface water (Yu et al. 2019). Numerous studies have illustrated that the naturally occurring DOM can significantly contribute to the photosensitive decomposition of environmental pollutants due to the strong oxidation or reduction capability of the photogenerated triplet 3DOM*, hydrate electrons (eaq), and reactive oxygen species (ROS), which are able to react with pollutants via direct oxidation, energy transfer, and/or electron transfer (Jiao et al. 2022). To verify whether the CNMs-FA can promote the photodegradation of pollutants in natural surface water, the photodegradation of selected contaminants, including EE2, DBP, and TC, was examined in the presence of CNMs-FA composites. Since the CNMs-FA composites are able to significantly adsorb the target organic contaminants, as discussed in Section 3.2, the photodegradation experiments were performed after pre-adsorption of organic contaminants on CNMs-FA for 12 h under dark.

As depicted in Figure 4, the different degradation behaviors were observed for three different organic contaminants after pre-adsorption processes, implying that the complex mechanisms are involved in the photodegradation of different pollutants on the surface of different CNMs-FA composites. For the control samples that only pollutant exists in aqueous solutions, no significant degradation of EE2 and DBP was observed (Figure 4(a) and 4(b)). Meanwhile, the TC decomposed about 35% in 8 h of exposure to sunlight (Figure 4(c)). These observations indicated that the direct photodegradation of EE2 and DBP in aqueous solutions is neglectable since there is no overlapping wavelength between the absorption bands of EE2 and DBP (<290 nm) and the light emission bands of simulated sunlight (320–780 nm) employed in this study, as presented in Supplementary Figures S2 and S9. However, the light absorption of TC extends to 410 nm (Supplementary Figure S9) and the overlapping light in the wavelength range of 320–410 nm may contribute to the direct photodegradation of TC under sunlight irradiation. A previous study also demonstrated that TC could be degraded directly under unlight or light irradiation from 290 to 800 nm (Song et al. 2022). In addition, Figure 3(a)–3(c) illustrates that all three studied contaminants undergo photosensitive degradation on the surface of three kinds of CNMs-FA composites although the degradation efficiencies were different. Similar to FA, CNMs-FA may play multiple roles in the photochemical degradation process of organic compounds under sunlight irradiation, including promoting photosensitive decomposition and inhibiting direct photolysis. As a photosensitizer, CNMs-FA might enhance the decomposition of organic compounds by producing the reactive species CNMs-3FA*, OH, and 1O2 (Yu et al. 2021; Jiao et al. 2022). As an inhibitor, the CNMs-FA could act as a light screening agent and quencher of OH radicals to prevent the photolysis of organic contaminants (Lindsey & Tarr 2000). To verify the possible contributions of the photogenerated species such as CNMs-3FA* and ROS to the photosensitive decomposition of organic contaminants, quenchers, including TMP for CNMs-3FA*, FFA for 1O2, and IPA for OH, were separately added to samples (pH = 7.0 ± 0.1) containing organic contaminants and CNMs-FA composites before the photolysis experiment. After 8 h of irradiation under simulated sunlight, the concentration changes of pollutants were compared with those without quenchers for the same samples to evaluate the effects of photogenerative reactive species on the decomposition of studied organic contaminants. As shown in Supplementary Figure S10, the addition of quenchers of OH and 1O2, including IPA and FFA, slightly inhibited the degradation of organic contaminants, suggesting that the decomposition of organic contaminants was partially driven by these photochemically produced reactive species in the presence of CNMs-FA under irradiation. In addition, due to the lower one-electron oxidation potential [Eo(TMP+•/TMP) = 1.2 VNHE] than the one-electron reduction potentials of most 3DOM* [Eo(3DOM*/DOM–•) = 1.4–1.9 VNHE], the TMP was selected as the electron transfer probe for the examination of the reaction mechanism between CNMs-3FA* and organic contaminants (Wan et al. 2021). According to the quenching experimental results, the addition of TMP markedly suppressed the photodecomposition of organic contaminants (Supplementary Figure S10), verifying that the CNMs-3FA* is mainly responsible for the photosensitive decomposition of organic contaminants via an electron transfer mechanism (Wan et al. 2021). Furthermore, the photosensitive degradation of all three organic contaminants in the presence of GO-FA (8.6–20%) was more stronger than those in the presence of SWCNT-FA (<5.6%) and MWCNT-FA (<6.2%). This observation was possibly attributed to the fact that except for the contribution of GO-3FA*, the excited state of GO* under light irradiation can not only directly react with dissolved O2 to generate 1O2 via energy transfer but also form the electron–hole pairs on the surface of GO. Then, the free electrons can combine with O2 to produce O2 and the holes may react with H2O or OH ions to form OH (Cao et al. 2021). Meanwhile, the photodegradation efficiency also depended on the chemical structure and properties of the organic pollutant itself. As illustrated in Figure 4(b), the SWCNT-FA and MWCNT-FA exhibited no effects on the DBP degradation and even the GO-FA slightly promoted the DBP decomposition about 5% exposing to sunlight irradiation for 8 h, suggesting that DBP is difficult to be degraded by the photo-induced reactive species comparing to EE2 and TC since the DBP consists of extremely stable benzene carboxylic group, which is in consistent with the previous studies on the photodegradation of organic pollutants in the presence of FA (Yu et al. 2019; Feng et al. 2022). Taken together, the experimental results reflected that the formed CNMs-FA composites are able to enhance the photosensitive degradation of organic contaminants due to the photo-induced reactive species such as ROS and CNMs-3FA* via direct oxidation, energy transfer, and electron transfer mechanisms. Of those, the electron transfer mechanism with CNMs-3FA* plays the most dominant role in the photosensitive decomposition of organic contaminants.
Figure 4

Pre-adsorption and photodegradation of organic contaminants in the presence of CNMs-FA composites at pH 7.0 ± 0.1 and 25 ± 1 °C. (a) EE2, (b) DBP, and (c) TC.

Figure 4

Pre-adsorption and photodegradation of organic contaminants in the presence of CNMs-FA composites at pH 7.0 ± 0.1 and 25 ± 1 °C. (a) EE2, (b) DBP, and (c) TC.

Close modal

The current study demonstrated that the environmental behavior of CNMs in the real aquatic environment mainly involves adsorbing naturally occurring DOM to form CNMs-DOM composites, which are able to adsorb and further promote the photodegradation of trace organic contaminants in surface water under sunlight irradiation. This study presented new data for the comprehensive elucidation of CNMs-DOM formation and their associated environmental impacts on the organic contaminants in environmental water. It would shed light on the investigation of the environmental behavior, environmental implications and applications of CNMs to meet the increasing demand for sustainable nanotechnology development.

We thank technicians from the Department of Chemistry at South-Central Minzu University for their operation of FTIR and BET experimental tests for this work. The work was jointly supported by the Research Funding (YZZ18018) and the Fundamental Research Funds for the Central Universities (CZT20017; KTZ20043), South-Central Minzu University.

All relevant data are included in the paper or its Supplementary Information.

The authors declare there is no conflict.

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Supplementary data