Micro- and nanoplastic-mediated phototransformation and bioaccessibility of fluorinated liquid crystal monomer in aquatic environments

Abstract

Micro- and nanoplastics are emerging pollutants that have attracted significant attention due to their potential to concentrate and transport coexisting organic pollutants in aquatic environments. Fluorinated liquid-crystal monomers (FLCMs) have also emerged as contaminants of concern, given their frequent occurrence, potential toxic effects, and propensity to co-occur with plastics in the environment. However, the influence of plastics on the environmental fate of FLCMs remains unclear yet. To address this knowledge gap, we investigated the accumulation of a key FLCM, 4-cyano-3-fluorophenyl 4-ethylbenzoate (CEB-F), on three common plastics, and examined the effects of nanoplastics on the phototransformation of CEB-F and its acute toxicity to Daphnia magna (D. magna). Our findings revealed that the adsorption capacity of CEB-F on different plastic materials followed the order: polystyrene (PS) < mixed cellulose ester (MCE) < polyamide (PA). The adsorption processes of CEB-F on the three plastics aligned more closely with the pseudo-first-order kinetic model and the Langmuir isotherm model, suggesting that the adsorption is primarily governed by physical diffusion. Theoretical calculations indicated that the adsorption of CEB-F on PS plastics is mainly driven by hydrophobic interactions. Additionally, PS nanoplastics (PSNPs) significantly enhanced the UV degradation of CEB-F, although the types of degradation intermediates did not change substantially, suggesting a limited impact on the degradation process and mechanism. Acute toxicity tests showed that PSNPs increased the toxicity of CEB-F to D. magna at lower concentrations, while the toxicity was reduced at higher concentrations. The obtained findings are of great significance to unraveling the plastic-mediated environmental fate and aquatic toxicity of FLCMs in natural waters.

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Environmental significance

Micro- and nanoplastic contamination has received a lot of attention because of its ability to concentrate and transfer organic contaminants in aquatic settings. Fluorinated liquid-crystal monomers (FLCMs) have also emerged as contaminants of concern due to their widespread prevalence, potential hazardous effects, and proclivity to coexist alongside plastics in the environment. Thus, it is critical to investigate the impact of micro- and nanoplastics on the environmental fate of FLCMs. In this investigation, we discovered that FLCMs were readily absorbed by common micro- and nanoplastics, primarily due to hydrophobic interactions. Nanoplastics considerably increased the phototransformation of FLCMs while playing complex roles in their toxicity to D. magna. The current findings are of tremendous significance to unraveling the plastic-mediated environmental destiny and aquatic toxicity of organic contaminants.

1. Introduction

Liquid crystal monomers (LCMs) are essential components of liquid crystal displays (LCDs), typically sealed between two glass substrates or other suitable materials to provide protection and support.¹ Numerous studies have reported the widespread application of LCMs in the electronic industry, including computers, televisions, and smartphones. Global production of LCMs increased from 500 tons in 2011 to 1300 tons in 2021, with continued growth expected.² Uncontrolled production and poor management of LCD products can lead to the release of abundant LCMs into the environment, particularly through e-waste recycling.³ Upon entry into the environments, released LCMs from LCDs are readily to undergo surface runoff, atmospheric deposition, and accumulation in global environmental matrices such as sediments,⁴ soil,⁵ and indoor dust.⁶ Additionally, due to their structural similarities with persistent organic pollutants like polybrominated diphenyl ethers (PBDEs) and organic phosphates (OPEs), LCMs can bioaccumulate in biota⁷ and potentially pose risks to human health and wildlife via food chains.^2,8 From a public health and biodiversity protection perspective, it is crucial to understand the environmental fate and potential ecological risks associated with LCMs.

Among the many LCM congeners synthesized and commercialized in current global markets, fluorinated liquid-crystal monomers (FLCMs) are particularly significant due to their high voltage retention, high resistivity, and good stability. FLCMs typically possess a biphenyl backbone structure with a range of chemical substitutions (i.e., alkyne, cycloethyl, cyano, ester, and fluorine substitutions), and have higher toxicity compared to other LCMs.⁹ This structure results in air–water partition coefficients (log K_aw) ranging from −5 to −1 and octanol–water partition coefficients (log K_ow) of the 27 FLCMs higher than 5, making FLCMs prone to widespread distribution and potential toxic risk to human health.¹⁰ However, the fate and toxic effects of FLCMs in the environment are still not fully understood, despite their ubiquitous presence, as indicated by previous research.

Micro- and nanoplastics, defined as plastic particles ranging from 1 μm to 5 mm and less than 1 μm, respectively, have garnered global concern due to their persistence in natural waters and toxic effects on biota, including human bodies.¹¹ These plastic particles can readily adsorb organic and inorganic contaminants, such as heavy metals, dyes, pesticides, drugs, and antibiotics, forming plastic–contaminant complexes through electrostatic forces, hydrophobic interaction, hydrogen bonding, π–π conjugation, and micropore filling.¹² Moreover, the universal presence of micro- and nanoplastics can alter the photochemical behavior of organic contaminants in aquatic environment.^13,14 For example, the phototransformation of atorvastatin (ATV) has been shown to positively correlate with the carbonyl index of photoaged polystyrene (PS) microplastics, as aged microplastics with more oxygen-containing functional groups (e.g., alcohols, aldehydes, carboxylic acids, and esters) can produce more reactive oxygen species (ROS) like singlet oxygen (¹O₂) and triplet-excited state PS (³PS*), thus enhancing ATV phototransformation rates in water.¹⁵

Furthermore, it was also found that micro- and nanoplastics can also alter the toxicity of organic contaminants in water. For instance, it has been reported that PS nanoplastics reduced the toxicity of polycyclic aromatic hydrocarbons (PAHs), as PAHs adhered to the plastic surface, decreasing their bioavailability and subsequent toxicity.¹⁶ Given these insights, we hypothesize that micro- and nanoplastics can significantly influence the fate and toxicity of organic compounds including LCMs in aquatic environments. However, the mechanisms underlying the effects of micro- and nanoplastics on the phototransformation and toxicity of LCMs in water remain largely unexplored.

This study focused on the micro- and nanoplastic-mediated phototransformation and toxic effects of FLCMs in aquatic environments, since both of which have emerged as significant environmental concerns.^7,17 Notably, various FLCMs have been detected in aquatic environments at concentrations of up to 1120 ng L⁻¹,¹⁸ highlighting their substantial potential to interact with widespread plastic particles. Consequently, it is essential to examine the possible interactions between these contaminants and assess their environmental fate and risks when coexisting in aquatic systems. Using 4-cyano-3-fluorophenyl 4-ethylbenzoate (CEB-F) as a model LCM compound,¹⁹ this study aimed to: (i) explore the carrier effect of micro- and nanoplastics on CEB-F in water; (ii) elucidate the micro- and nanoplastic-mediated phototransformation of CEB-F, and (iii) investigate the bioaccessibility and toxic effects of the plastic–CEB-F complex. The findings will contribute to our understanding of the fate and potential ecological risks of CEB-F associated with plastic particles in natural waters.

2. Materials and methods

2.1. Chemicals and reagents

Mixed cellulose ester (MCE), polyamide (PA), and polystyrene (PS) plastic pellets used in this experiment were obtained from common plastic products in life. Nano-sized polystyrene (PSNPs, 0.05–0.1 μm) was purchased from Macklin Biochemical Technology Co., Ltd. (Shanghai, China). 4-Cyano-3-fluorophenyl-4-ethylbenzoate (CEB-F, >97% purity) was purchased from Bide Pharmatech Ltd. (Shanghai, China). Sodium sulfite (Na₂SO₄, ≥99% purity), sodium nitrate (NaNO₃, ≥99% purity), sodium bicarbonate (NaHCO₃, ≥99.8% purity), and sodium chloride (NaCl, ≥99.5% purity) of analytical grade were purchased from Shanghai Titanchem Co., Ltd (Shanghai, China). HPLC-grade acetic acid and acetonitrile were supplied by Aladdin (Shanghai, China). Deionized (DI) water prepared from a Millipore Milli-Q system (>18.2 MΩ cm⁻¹) was used for solution preparation. All other chemicals and reagents were of analytical grade. A stock solution of CEB-F at 5.0 mM was prepared in acetonitrile and stored at 4 °C for further use.

2.2. Characterization of micro- and nanoplastics

The surface morphology of MCE, PA, and PS plastics was examined through scanning electron microscopy (SEM, Hitachi S4800, Japan). The chemical structures of the four types of plastics were analyzed using Fourier-transform infrared spectroscopy (FT-IR, Thermo Nicolet 6700) via a pressurization method. The surface chemical composition and elemental binding energies were determined through X-ray photoelectron spectroscopy (XPS, Thermo VG ESCALAB 250, USA) with monochromatic Al Kα radiation. Zeta potential measurements at different pH levels and at 25 °C were determined using a Malvern Zetasizer Nano series Nano-ZS (MA, USA), which were computed using Zetasizer Software (Version 7.10, MA, USA) based on Henry's function and the Smoluchowski approximation. Additionally, the hydrophobic properties, represented by the water contact angle, were measured with a dynamic contact angle instrument (JY-82B Kruss DSA, Dataphysics OCA20, Shimadzu, Japan).

2.3. Batch adsorption experiments

Regarding the adsorption kinetics of different types of plastics (i.e., MCE, PA, and PS) to CEB-F in water, we exposed 10 mg L⁻¹ of plastics and 10 μM of CEB-F and incubated at 125 rpm in a shaker that operated at 25 °C. Then, 1 mL of the reaction solution was removed at 5, 10, 30, 45, 60, 90, 120, 180, 240, and 360 min, respectively, and determined the residue CEB-F within the reaction system. In addition, to illustrate the impact of water constituents (pH, anion, and salinity) on the sorption of PSNPs to CEB-F in water, we mixed 10 mg L⁻¹ of PSNPs with different pH (3, 4, 5, 6, 7, 8, and 9, adjusted using 0.01 mol L⁻¹ NaOH/HCl), 10 mM ions (Cl⁻, SO₄²⁻, NO₃⁻, Mg²⁺, and Ca²⁺) and salinity (0.10%, 0.50%, 1.00%, 2.00%, and 3.00%).

The sorption amount of plastics to CEB-F can be calculated as follows (eqn (1)),

where the q_e represents the equilibrium sorption amounts of plastics to CEB-F. C₀ and C_e represent the concentration of CEB-F (μM) before sorption experiments and equilibrium sorption time, respectively. V (mL) is the volume of the reaction solution, and m (g) is the quality of the adsorbent in the reaction system.

Pseudo-first- and second-order kinetic, as well as intraparticle diffusion models were applied for the sorption kinetic analysis of CEB-F to different types of plastics, with the equations as follows (eqn (2) to (4)),

q_t = q_e(1 − e^−k₁t) (2)

q_t = k_it^0.5 + c (4)

where t (min) is the reaction time. q_t is the sorption amount of plastic to CEB-F at time t. q_e is the equilibrium sorption amount of plastic to CEB-F. k₁ and k₂ (μmol g⁻¹ min⁻¹) are of the correlation rate constants in pseudo-first-order and second-order models, respectively. k_i (μmol g⁻¹ min^−0.5) is the diffusion rate constant. c is the boundary-layer thickness.

Adsorption isotherm of plastics to CEB-F was conducted by mixing 10 mg L⁻¹ of PSNPs with 5, 10, 20, 40, and 80 μM of CEB-F and shaken for 240 min at 125 rpm and 25 °C. Then 1 mL of aquatic solution was removed from the reaction system and deterring residue CEB-F concentration. Langmuir (eqn (5)) and Freundlich (eqn (6)) models were utilized fitting the sorption amount of plastics to CEB-F in water, with the specific calculation strategies as follows (eqn (5) and (6)),

where q_max (μmol g⁻¹) is the maximum sorption amount of plastic to CEB-F in water. k_L (L μmol⁻¹), k_F ((μmol g⁻¹) (μmol L⁻¹)^−1/n), and 1/n are of the Langmuir adsorption constant, Freundlich adsorption constant, and the Freundlich parameter, respectively.

2.4. Phototransformation experiments

The phototransformation of CEB-F in water was conducted using a photochemical reactor (XPA-7, Xujiang, Nanjing) which coupled with an 800 W xenon lamp (UV intensity: 16.76 mW cm⁻², wavelength range 200–1000 nm) and a 300 W medium pressure mercury lamp (UV intensity at 5.0 mW cm⁻², with the wavelength range of 200–400 nm) to simulate sunlight and ultraviolet light irradiation, respectively. The reaction temperature was maintained at 25 °C using a cyclical cooling water system. PBS buffer was used to maintain the stability of the solution's pH. At 0, 1, 2, 4, 6, 10, 15, and 30 min of UV irradiation, 1 mL of the reaction solution was sampled and mixed with 2 mL of methanol to desorb any chemicals adhered to PSNPs. The mixture was then passed through a 0.45 μm membrane filter and analyzed using high-performance liquid chromatography (HPLC).

The quantification of CEB-F was achieved by a Shimadzu LC-20AT HPLC (Shimadzu, Japan) equipped with a UV detector and a 4.6 × 250 mm Agilent Zorbax SB-C18 reverse-phase column (Agilent, USA) maintained at 35 °C. The UV detector wavelength was set at 254 nm, and the injection volume was set at 20 μL. The isocratic mobile phase was made up of 80% acetonitrile and 20% DI water (with 0.3% acetic acid) with a flow rate of 1.0 mL min⁻¹. The phototransformation intermediates of CEB-F were identified using an ultra-high-performance liquid chromatography-tandem triple quadrupole mass spectrometer (UPLC-Q-Exactive Orbitrap-MS, Thermo, Bremen, Germany) with a Hypersil GOLD C18 column (100 × 2.1 mm, 1.9 μm). The mobile phase consisted of 0.1% formic acid in deionized water and acetonitrile (80 : 20, v/v), with a flow rate of 0.3 mL min⁻¹. The injection volume and the column temperature were maintained at 10 μL and 35 °C, respectively.

2.5. Computational methods

The interaction and binding of CEB-F with PS plastics were calculated using the Forcite module in the Materials Studio software. The PS molecule was modeled with a polymerization degree of 20 repeating units, consisting of 160 carbon atoms and 162 hydrogen atoms. The initial model structure was optimized using the Smart algorithm, followed by molecular dynamics simulations. The simulations were carried out under NPT conditions at 298 K and 1 atm pressure, with a total equilibrium time of 100 ps. The COMPASS II force field was applied, using the Berendsen thermostat for temperature control and the Berendsen barostat for pressure control,²⁰ with a time step of 1 fs. To ensure accurate simulation results, the system was considered equilibrated when temperature and energy fluctuations during the final NPT molecular dynamics simulation were within the range of 5% to 10%.²¹

2.6. Daphnia magna exposure experiments

The bioaccessibility and toxicity of CEB-F mediated by plastics to aquatic organisms were assessed using Daphnia magna as an indicator species, following the standard protocol outlined in OECD Guideline No. 202. Third-brood neonates (<24 h old) were used for the toxicity tests. Prior to the experiments, Daphnia magna was cultured at 24 °C with a daily light/dark cycle of 8 h/16 h. During the bioaccessibility and toxicity experiments, 10 daphnids were exposed to 20 mL of CEB-F at concentrations of 0, 0.2, 0.5, 1.0, 1.5, and 5 μM, with 0, 2 mg L⁻¹, and 5 mg L⁻¹ PSNPs, respectively. The survival rate was recorded after 96 hours of incubation without feeding. Catalase (CAT) and acetylcholinesterase (AChE) activities (expressed as U mg⁻¹ protein) were measured using commercial kits from Beijing Boxbio Science & Technology Co., Ltd., following the manufacturer's instructions. Protein content was determined by the Bradford method.^22,23

2.7. Statistical analysis

Each test was repeated at least three times, and finally presented as the mean ± standard deviation. Statistical analyses related to CEB-F sorption kinetics and isotherms were conducted using Origin 2018 software. Differences were statistically analyzed by one-way analysis of variance (ANOVA) followed by the LSD test for multiple comparisons (p < 0.05, n = 3) using SPSS 18.0 (IBM Company).

3. Results and discussion

3.1. Characterization of micro- and nanoplastics

The surface morphology of PS, MCE, and PA microplastics (MP) and PS nanoplastics (NP) was analyzed using SEM, as shown in Fig. 1a. The purchased PSNPs are spherical particles with smooth surfaces and an average size of approximately 80 nm. Both PSMPs and PAMPs exhibited smooth bulk structures, while MCEMPs displayed a distinctive cross-linked porous structure. The size distributions for these plastics ranged from 1 to 300 μm. The surface elemental compositions of the four plastics were characterized via XPS (Fig. 1b). The survey spectra revealed the presence of C and O on the surfaces of PSNPs and PSMPs, whereas MCEMPs and PAMPs also contained N (Fig. 1b). High-resolution XPS spectra of C 1s, O 1s, and N 1s for all four plastics are shown in Fig. S1.† PSNPs and PSMPs exhibited three distinct C 1s peaks at 284.2, 285.8, and 291.6 eV (Fig. S1a and b†), corresponding to aromatic C–H, aliphatic C–H, and π–π* transitions, respectively.^24,25 MCEMPs displayed four C 1s peaks at 284.6, 286.2, 287.0, and 288.1 eV (Fig. S1c†), attributed to aromatic C–H, aliphatic C–H, C–O/C–N, and C=O groups.²⁶ Similarly, PAMPs showed three peaks at 284.4, 285.8, and 287.8 eV (Fig. S1d†), representing aliphatic C–H, C–N, and C=O groups.²⁷ The O 1s peak fitting for four plastics is similar, showing all have –OH, –C–O, and –O–C=O groups. MCEMPs showed three N 1s peaks at 398.5, 407.5, and 408.2 eV, corresponding to C–N, –NO₂, and –ONO₂ groups.²⁶ For PAMPs, two peaks at 399.1 and 399.8 eV were attributed to N–H and C–N groups, respectively (Fig. S1d†).

Fig. 1 Characterization of the four plastic particles. (a) SEM images, (b) XPS spectra, (c) FT-IR spectra, (d) zeta potential values under various pH, (e) water contact angle, and (f) the average water contact angle for left and right (n = 3). Differences were statistically analyzed by one-way analysis of variance (ANOVA) followed by the LSD test for multiple comparisons using SPSS 18.0 (IBM Company). In panel (f), means marked with different letters indicate significant differences between treatments (p < 0.05, n = 3).

Fig. 1c depicts the functional groups distribution of four plastics as characterized by FT-IR. PSNP and PSMP include 3440 (–OH), 3082–2849 (–CH₂), 1740 (C=O), 1600–1450 (C=C), 1372 (C=O), 1069 (–C–O) cm⁻¹, respectively. PAMPs has the peaks at 3306 (–OH, –NH), 3081–2859 (–CH₂), 1637–1461 (C=C), 1371 (C=O) cm⁻¹, while MCEMP possess peaks at 3400 (–OH), 1748 (O–C=O), 1640 (C=O), 1373 (C=O, –NO₂), 1275 (C–N), 1069 (–C–O) cm⁻¹, respectively.^24–28 Zeta potential measurements in Fig. 1d revealed that both PSNP and PSMP carried a negative charge across a pH range of 2.0 to 10.0. In contrast, MCEMPs and PAMPs exhibited positive surface charges under acidic conditions, with zero-point charge (pH_ZPC) values of 4.18 and 4.35, respectively.

Furthermore, water contact angle was measured to evaluate the hydrophobicity of the four plastics,¹⁵ as shown in Fig. 1e and calculated in Fig. 1f. Water contact angles of PSNP, PSMP, MCEMP, and PAMP were 104.7°, 116.6°, 101.5°, and 105.7°, respectively, indicating the hydrophobicity of these four plastics followed the order of PSMP > PSNP ≈ PAMP > MCEMP.

3.2. Carrier effect of micro- and nano-plastics to CEB-F in aqueous solution

The adsorption behaviors of CEB-F by commonly present plastics, specifically PS, PA, and MCE were compared. As shown in Fig. 2a and Table S1,† the pseudo-first-order kinetic model provided higher correlation coefficients than the pseudo-second-order kinetic model for all three types of plastics, indicating that the adsorption of CEB-F was primarily driven by physical adsorption. The results from the intra-particle diffusion model (Fig. 2b) revealed that the linear fits did not pass through the origin for any of the three materials, suggesting that both surface adsorption and intra-particle diffusion played roles in the actual adsorption process of CEB-F on these plastics.

Fig. 2 Carrier effect of mixed cellulose ester (MCE), polyamide (PA), and polystyrene (PS) microplastics (MPs), as well as PS nanoplastics (PSNPs) to CEB-F in aquatic solutions. (a) CEB-F adsorption kinetic curves on various plastics, and (b) the intraparticle diffusion model fitting results.

Initially, approximately 80% of CEB-F was adsorbed onto the plastics, likely due to multiphase adsorption interactions such as hydrophobic interactions, covalent bonds, and van der Waals forces, which occupied the external activation sites. During this stage, the significant thickness of the plastic boundary layer highlighted the importance of surface adsorption in the process.^29–31 Additionally, the q_t and t^0.5 did not pass through the origin, indicating that the rate-limiting step was influenced by both membrane diffusion and intra-particle diffusion.³⁰ In the second stage, only 10% of CEB-F continued to adsorb on the plastic surface. This phase involved slow diffusion from the liquid film to the microporous surface. However, due to the relatively small diameter of the micropores and the large molecular size of CEB-F, the intra-particle diffusion rate was markedly slower than in the initial stage.^29,30

The pseudo-first-order model's simulated sorption rate constants for CEB-F on the three plastic MPs (MCE, PA, PS) in water were 0.0193, 0.0062, and 0.6609 min⁻¹, respectively. The pseudo-second-order model's simulated rate constants for three plastics were 0.0017, 0.0006, and 0.0914 μmol g⁻¹ min⁻¹, respectively. These results indicate that the sorption rate of CEB-F on PS MPs in water was higher than on MCE and PA MPs. However, according to the pseudo-first-order simulation results, the equilibrium sorption amounts of CEB-F on the three plastics were ranked as follows: PS (14.1503 μmol g⁻¹) > MCE (10.8453 μmol g⁻¹) > PA (6.9769 μmol g⁻¹). These findings suggest that PS plastics have a high potential to adsorb CEB-F onto their surface in water, forming plastic–contaminant complexes.

Previous studies have estimated and demonstrated that the octanol–water partition coefficient (log K_ow) values of most LCMs exceed 5, highlighting their high hydrophobicity. This property suggests a strong tendency for these compounds to adsorb onto surfaces with similar hydrophobic characteristics, such as sediments and fatty tissues in organisms.^4,32,33 Furthermore, prior research has shown that the adsorption of hydrophobic pollutants, such as tylosin, onto microplastics like PE, PP, PS, or PVC is primarily driven by hydrophobic interactions.³⁴ The higher adsorption of CEB-F onto PSMP can be explained by the fact that PSMP is more hydrophobic than PAMP and MCEMP, according to measurements of water contact angles (Fig. 1e and f). Additionally, the SEM image in Fig. 1a showed that MCEMPs possess a porous structure, which facilitates the capture of CEB-F, further contributing to its adsorption potential.³⁵ In this study, the adsorption capacity of the three typical plastics for CEB-F reached the μmol g⁻¹ level under simulated conditions, aligning with most studies on persistent organic pollutants (POPs).³⁶ Despite environmental concentrations being lower than the modeled concentrations, plastics were still demonstrated to be significant carriers of FLCMs.

Further comparison between the adsorption of CEB-F by PSMPs and PSNPs revealed that, although the simulated pseudo-first-order sorption rate constant for PSNPs (0.0570 min⁻¹) was much lower than that for PSMPs (0.6609 min⁻¹), the equilibrium sorption amount for PSNPs (589.41 μmol g⁻¹) was significantly higher than that for PSMPs (14.15 μmol g⁻¹). This indicates that nanosized plastics have a high potential to adsorb CEB-F onto their surface in water, forming plastic–contaminant complexes.

The sorption affinity of PS MPs and NPs for CEB-F in water was explored through adsorption isotherms constructed at different CEB-F concentrations. As shown in Fig. S2,† the sorption amount of CEB-F by both PS MPs and NPs increased continuously with rising CEB-F concentrations in water. The fitting results of the Langmuir, Freundlich, and Henry isotherm models for the adsorption of CEB-F by plastics (Table S2†) indicated that the Langmuir isotherm model best described the adsorption process for both PS MPs (R² = 0.9998) and PS NPs (R² = 0.9851), compared to the Freundlich and Henry isotherm models. This suggests that surface adsorption is the primary mechanism for CEB-F adsorption by PS MPs and NPs in water. Overall, the sorption kinetics and isotherm model results emphasize that PS MPs and NPs are crucial carriers for CEB-F in aquatic environments.

3.3. Influence of the initial pH, ions, and salinity on CEB-F adsorption by nanoplastics

By using PSNPs as the represent, we further examined the influence of nature water constituents and properties (e.g., pH, ions, and salinity) on CEB-F adsorption by nanoplastics. pH is a crucial environmental factor that affects both the morphology of the adsorbent and its surface physicochemical properties.³⁷ As shown in Fig. 3a, the adsorption capacity of PSNPs for CEB-F initially increases and then decreases as the pH rises from 3.0 (153 μmol g⁻¹) to 9.0 (324 μmol g⁻¹), with the maximum adsorption capacity observed at pH 8.0 (368 μmol g⁻¹). Overall, the sorption affinity of CEB-F to PSNPs is higher in alkaline conditions (pH > 7) compared to acidic conditions (pH < 7). Zeta potential measurements in Fig. 1d indicate that PSNPs were negatively charged across the pH range of 2.0 to 10.0, with the zeta potential values gradually decreasing as pH increased. Under alkaline conditions, the surface of PSNPs exhibited a greater negative charge, which may enhance electronic interactions such as the π–π electron donor–acceptor effect and hydrogen bonding between PSNPs and CEB-F.^38–40 Additionally, as the pH increased, the stronger negative charge on PSNPs likely reduced homoaggregation due to increased electrostatic repulsion. This reduction in aggregation is favorable for improving CEB-F adsorption onto PSNPs.⁴¹

Fig. 3 Effect of (a) pH, (b) ion species, (c) salinity, and (d) water matrix on the adsorption capacity of CEB-F on PSNPs. Differences were statistically analyzed by one-way analysis of variance (ANOVA) followed by the LSD test for multiple comparisons using SPSS 18.0 (IBM Company). In each figure, means marked with different letters indicate significant differences between treatments (p < 0.05, n = 3).

To unravel the effect of anions and cations in natural water on the adhesion of CEB-F to plastics, we compared the sorption capacity of PSNPs to CEB-F under 10 mM of Cl⁻, SO₄²⁻, NO₃⁻, Mg²⁺, and Ca²⁺, respectively. In Fig. 3b, in comparison with blank controls (643 μmol g⁻¹), the presence of Cl⁻ and NO₃⁻ (415 and 307 μmol g⁻¹) obviously inhibited the sorption capacity of PSNPs to CEB-F while the presence of SO₄²⁻, Mg²⁺, and Ca²⁺ (637, 613, and 586 μmol g⁻¹) displayed exiguous effect on the sorption capacity of PSNPs to CEB-F.

In addition to pH and ions, solution salinity also affects the sorption affinity of PSNPs to CEB-F in water. As illustrated in Fig. 3c, the sorption capacity of PSNPs to CEB-F decreases continuously from 643 μmol g⁻¹ to 215 μmol g⁻¹ as the solution salinity increases from 0 to 3.5%. The mechanism behind salinity's impact on PSNPs adsorption is due to the increased solution salinity causing PSNP particles to aggregate, thereby reducing the specific surface area and the electric double layer of PSNPs. This aggregation decreases the available sorption sites for PSNPs CEB-F.⁴² For example, it has been observed that carboxylated polystyrene MPs with a particle size of 30 nm rapidly aggregate to sizes above 1000 nm in less than 30 minutes in seawater.⁴³ Moreover, the presence of NaCl readily increases the viscosity and density of the solution, thereby inhibiting the mass transfer from the water phase to the solid phase.⁴⁴ Consequently, higher aquatic salinity reduces the surface charge of PSNPs, diminishes the electrostatic interaction between CEB-F and PSNPs, and ultimately weakens the adsorption capacity of PSNPs for CEB-F in water. For better understanding the carrier effect of plastics particle on FLCMs in the aquatic environment, the adsorption ability of PSNPs to CEB-F in different natural water matrices (tap, river, and lake water, primary properties included in Table S3†) was examined. As shown in Fig. 3d, the adsorption amounts of CEB-F were not influenced by the mater matrix, confirming again that PSNPs are important carriers for FLCMs in natural surface water.

3.4. Sorption mechanism of PSNPs to CEB-F in water

To elucidate the sorption mechanism of PSNPs for CEB-F, FTIR was performed to assess changes in surface functional groups before and after adsorption. As depicted in Fig. S3,† pure CEB-F exhibits characteristic peaks at 1748 cm⁻¹ (aryl C=C ring vibrations), 1640 cm⁻¹ (–C=O), 1275 cm⁻¹ (C–F), 1057 cm⁻¹ (C–O), and 839 cm⁻¹ (C–H vibrations) cm⁻¹, respectively. Pure PSNPs, on the other hand, display peaks at 3443 cm⁻¹ (–OH), 3045–2922 cm⁻¹ (–CH₂), 1740 cm⁻¹ (C=O), 1605–1452 cm⁻¹ (C=C), 1372 cm⁻¹ (C=O), and 1069 cm⁻¹ (–C–O).⁴⁵ After adoption of CEB-F, an increase in the characteristic peak at 1371 cm⁻¹, was observed, suggesting the formation of C–N bonds,⁴⁶ and an increase at 1067 cm⁻¹, indicates the formation of C–F bonds. Additionally, the characteristic peak of PSNPs at 1605 cm⁻¹ shifted to 1621 cm⁻¹, possibly due to physical overlap resulting from CEB-F adsorption. However, due to the limited adsorption of CEB-F, the changes in these characteristic peaks were subtle. The surface interactions between PSNPs and CEB-F were further examined via high-resolution C 1s, O 1s, and F 1s XPS spectra of pure PSNPs, pure CEB-F, and the PSNP–CEB-F complex. The survey spectra confirmed the presence of C and O on PSNPs, while CEB-F exhibited C, O, N, and F (Fig. S4†). After adsorption, a weak F peak appeared, confirming the adsorption of CEB-F onto PSNPs. High-resolution XPS spectra of C 1s, O 1s, and F 1s are shown in Fig. S5.† PSNPs displayed three distinct C 1s peaks at 284.2, 285.8, and 291.6 eV, corresponding to aromatic C–H, aliphatic C–H, and π–π* transitions.^24,25 Following adsorption, the π–π* transition peak at 291.6 eV shifted to a lower binding energy of 291.1 eV, indicating π–π electron donor–acceptor (EDA) interactions between the benzene rings of CEB-F and PSNPs.^38–40 For O 1s spectra, the –OH, –C–O, and –O–C=O groups of pure PSNPs were observed at 531.2, 532.0, and 532.9 eV, while for pure CEB-F, they were located at 531.8, 532.6, and 533.5 eV, respectively. After adsorption, these peaks for the PSNP–CEB-F complex shifted slightly to 531.6 eV (–OH), 532.5 eV (–C–O), and 533.2 eV (–O–C=O), further confirming the interactions between CEB-F and the PSNP surface. The F 1s spectrum of CEB-F displayed a distinct C–F peak at 687.5 eV, which was absent in pure PSNPs. After adsorption, a weak C–F peak emerged at 687.8 eV, suggesting the possible formation of weak halogen bonds between PSNPs and CEB-F.³⁴

To further explore the sorption mechanism, molecular dynamic simulations of the interaction between PSNPs and CEB-F were conducted. The simulation ensured statistical equilibrium accuracy, with temperature (Fig. S6a†) and energy (Fig. S6b†) fluctuations maintained within 5–10%, confirming system equilibrium. Fig. 4 illustrates the molecular dynamic binding process between CEB-F and PSNPs. The simulation revealed that after 10 ps of interaction, CEB-F spontaneously moved to the PSNP surface, with a binding energy of −48.4075 kcal mol⁻¹ (Table S4†). During the adsorption process, the PS polymer gradually bent and encapsulated CEB-F at 100 ps. The forces governing the PSNP–CEB-F interaction were primarily attributed to micropore filling and hydrophobic interactions, due to the hydrolytic nature of the carbon-based PSNP and CEB-F.

Fig. 4 The molecular dynamic binding process between CEB-F and PSNPs model.

3.5. Plastic-mediated phototransformation of CEB-F in aquatic solutions

Since PSNPs showed the highest adsorption capacity toward CEB-F, we chose them as the subject of our investigation in order to better explain the plastic-mediated environmental behavior of CEB-F in natural waters. This allowed us to explore in greater detail the role of plastics in the phototransformation of CEB-F. As shown in Fig. 5a, 10 μM of CEB-F can be degraded under the existence of 0, 2, 5, and 10 mg L⁻¹ PSNPs after 30 min of UV irradiation. The phototransformation of CEB-F with the presence of PSNPs can be well-described by a pseudo-first-order kinetics model with all R² > 0.98. The pseudo-first-order rate constant (k_obs) of the UV irradiation of CEB-F in the presence of PSNPs varying from 0 to 10 mg L⁻¹ was calculated and is presented Fig. 5b. The phototransformation rate of CEB-F exhibited a positive correlation with the concentration of PSNPs (R² = 0.87, p < 0.0001), indicating that PSNPs significantly promote the phototransformation of CEB-F in aquatic environments. As indicted in previous studies and in Fig. S7,† PSNPs exhibit strong ultraviolet absorption, which enables the photochemical generation of reactive species, such as free radicals and singlet oxygen. In addition, PSNPs can act as photosensitizers, facilitating the photochemical transformation of pollutants through energy transfer or electron transfer mechanisms.^15,47 These factors likely contribute to the enhanced phototransformation of CEB-F observed in the presence of PSNPs.

Fig. 5 Effect of PSNPs on the phototransformation behaviors of CEB-F. Phototransformation kinetics (a) and the obtained pseudo-first-order rate constants (k_obs) (b) of CEB-F under the presence of 0, 2, 5, and 10 mg L⁻¹ of PSNPs in water; HPLC chromatograms of sole CEB-F (c) and in the presence of 10 mg L⁻¹ of PSNPs (d) after degradation under UV irradiation. Differences were statistically analyzed by one-way analysis of variance (ANOVA) followed by the LSD test for multiple comparisons (p < 0.05, n = 3) using SPSS 18.0 (IBM Company).

Furthermore, the presence of PSNPs facilitated the formation of CEB-F degradation by-products, as identified by LC-MS analysis. Four primary phototransformation products (P1–P4) were detected (Fig. 5c and S8†). These products are formed through reactive oxygen species attack, hydroxylation, and C–O/C–F bond cleavage.^19,48,49 The distribution spectrum of CEB-F phototransformation products in the presence of PSNPs closely resembled that of the control group without PSNPs (Fig. 5d), indicating that the phototransformation pathways and mechanisms remain largely unchanged by the presence of PSNPs.

The temporal evolution of the four phototransformation products was tracked by recording their peak areas in the HPLC spectrum over the reaction time course (Fig. S9†). P1 showed maximum accumulation at 2 minutes, followed by a rapid decline and complete disappearance after 10 minutes, suggesting high reactivity and rapid degradation. The peak areas for P2, P3, and P4 initially increased rapidly within the first 10 minutes and then gradually decreased, indicating their propensity for further photolysis.

While the degradation rate of CEB-F increased in the presence of PSNPs, the phototransformation of intermediates was also affected. Notably, the generation and subsequent elimination of P1 were slower in the presence of PSNPs. For example, P1 vanished completely within 30 minutes in the absence of PSNPs but was still detectable in the presence of PSNPs. Conversely, the presence of PSNPs accelerated the elimination of P2. It is well-known that plastic particles, including PSNPs, exhibit photochemical reactivity and can degrade under long-term light irradiation, releasing dissolved organic matter and additives in the process.^50–53 Consequently, the dynamic interactions between PSNPs and CEB-F as well as its intermediates significantly influence the rates and pathways of phototransformation, which should be highlighted in future investigations.

3.6. Toxicity change

To evaluate the toxicity of CEB-F and plastic mixtures during photochemical reactions, we used Daphnia magna (D. magna) as a model aquatic organism. The survival data after 24 hours of exposure are shown in Fig. 6a. CEB-F exhibited significant acute toxicity to D. magna, with survival rates dropping to 50%, 25%, and 0% at concentrations of 1.0, 1.5, and 2.0 μM, respectively. PSNPs also demonstrated toxicity to D. magna; the survival rate decreased to 67% in the presence of 5 mg L⁻¹ of PSNPs, whereas PSMPs at the same concentration did not significantly affect survival. Interestingly, PSMPs reduced the acute toxicity of CEB-F to D. magna. With 5 mg L⁻¹ of PSMPs, the survival rate of D. magna increased from 55% to 87.8%, 25% to 69.5%, 0% to 32.8%, and 0% to 12.9% when exposed to 1.0, 1.5, 2.0, and 5.0 μM of CEB-F, respectively. This reduction in toxicity is likely due to the adsorption of CEB-F by PSMPs, while the reduced of PSMPs by D. magna due to its larger particle size compared to PSNPs, which hence reduced the intake of CEB-F by D. magna.

Fig. 6 (a) Toxic effect of PSMPs and PSNPs on the survival of Daphnia magna at different CEB-F concentration intervals; (b) effect of PSMPs and PSNPs on the CAT and AChE enzyme activity under at different concentration of CEB-F. Differences were statistically analyzed by one-way analysis of variance (ANOVA) followed by the LSD test for multiple comparisons. In each figure, means marked with different letters indicate significant differences between treatments (p < 0.05, n = 3). A t-test was used to compare the means of the control group with each experimental group, a two-tailed p < 0.05 (*) and p < 0.01 (**) was considered statistically significant. All analyses were conducted using SPSS 18.0 (IBM Company).

However, the interaction between PSNPs and CEB-F presented a more complex toxicity profile. The presence of PSNPs at 5 mg L⁻¹ significantly reduced the survival of D. magna from 100% to 52.3% and from 55% to 43.2% at 0.2 and 1.0 μM of CEB-F, respectively. Due to their small size, nanoplastics can accumulate in D. magna, disrupt mitochondrial metabolism, induce intracellular oxidative stress, and compromise cell membrane integrity.^54–58 The accumulation of CEB-F on nanoplastics might extend the duration of CEB-F exposure within D. magna, potentially enhancing its toxicity. However, at higher concentrations of CEB-F (1.5, 2.0, and 5.0 μM), PSNPs alleviated the acute toxicity to D. magna. Our previous investigation found that CEB-F had significant acute toxicity on D. magna, with LC50 values of 1.0 μM after 24 hours.¹⁸ As a result, as CEB-F concentration increased, it began to dominate D. magna toxicity, whereas adsorption of CEB-F by PSNPs alleviated the acute toxicity to D. magna due to reduced bioavailability and bioaccessibility. At lower concentrations, PSNPs can enhance the toxicity of CEB-F by promoting its absorption and bioaccumulation.^54,57,58 On the other hand, at higher concentrations, PSNPs may mitigate CEB-F toxicity by forming a protective barrier and triggering defense mechanisms in organisms.^59,60

We further analyzed the activities of the enzymes catalase (CAT) and acetylcholinesterase (AChE) to clarify the role of PSMPs and PSNPs in mediating CEB-F aquatic toxicity. As shown in Fig. 6b, CAT activity significantly decreased with increasing concentrations of CEB-F from 0 to 1 μM. This reduction in CAT enzymatic activity indicates that elevated CEB-F exposure facilitates its penetration into D. magna, heightening oxidative stress and leading to an overproduction of H₂O₂ and HO˙. These reactive species exceed CAT's capacity to neutralize them, resulting in cellular damage.⁵⁹ Interestingly, the presence of PSNPs mitigated the inhibition of CAT activity by CEB-F, thereby protecting D. magna from cellular damage caused by free radicals.⁶⁰ Similarly, CEB-F exposure at 0.2 and 1.0 μM reduced AChE activity, indicating the neurotoxicity of CEB-F to D. magna.^61,62 However, the presence of plastics, both PSMPs and PSNPs, increased AChE levels in the mixtures, suggesting a reduction in the neurological effects of CEB-F on D. magna. These results indicate that the adsorption of CEB-F by plastics may reduce its interaction with CAT and AChE, thus altering the overall toxicity profile when CEB-F coexists with plastics.⁶³

4. Conclusions

This study explored the adsorption behavior, influencing factors, and phototransformation of CEB-F in the presence of micro- and nanoplastics (MPs and NPs) composed of polystyrene (PS), polyamide (PA), and mixed cellulose ester (MCE). The findings revealed that PS plastics, particularly PSNPs, have a high potential for adsorbing CEB-F due to their larger specific surface area and stronger hydrophobic interactions. The adsorption kinetics were best described by the pseudo-first-order model, and the adsorption process was primarily physical, involving surface adsorption and intra-particle diffusion. FTIR analysis and chemical calculations indicated that the adsorption mechanism involved hydrophobic interactions rather than chemical bonding. The presence of PSNPs was found to enhance the phototransformation of CEB-F under UV irradiation, increasing both the rate and extent of degradation. The phototransformation pathways remained consistent, although PSNPs altered the distribution and persistence of degradation by-products. Toxicity assessments using Daphnia magna demonstrated that CEB-F and PSNPs possess acute toxicity, with PSNPs exacerbating the toxic effects at lower CEB-F concentrations. However, PSMPs and PSNPs mitigated the toxicity at higher concentrations by reducing the bioavailability of CEB-F. Enzyme assays showed that PSNPs alleviated the inhibitory effects of CEB-F on CAT and AChE activities, suggesting a protective role against oxidative and neurotoxic damage. Considering the diverse interactions between plastics and CEB-F, as well as its degradation products, further evaluation is required to assess how the presence of plastic particles influences the toxicity of these products. Overall, the study highlights the significant role of MPs and NPs in modulating the environmental fate and toxicity of CEB-F, emphasizing the need for further research on the environmental impact of plastic–contaminant interactions.

Data availability

The authors confirm that the data supporting the findings of this study are available within the article and its ESI.†

Author contributions

All authors contributed to the conduct of the research and agreed to submit the manuscript. Y. F.: conceptualization, writing – original draft, visualization, writing – review and editing. J. W.: methodology, visualization, review and editing; W. L.: visualization, methodology, and formal analysis. W. Y.: methodology and formal analysis. D. G. visualization, methodology, and formal analysis. X. W.: writing – review, editing and visualization; Z. W.: investigation, writing and editing. R. W. S. L.: writing – review and editing.

Conflicts of interest

The authors declare no competing financial interest.

Acknowledgements

This research was financially supported by the National Natural Science Foundation of China (21707019), the Natural Science Foundation of Guangdong Province (2021A1515010019). X. W. acknowledges funding support from the National Natural Science Foundation of China (22376097, 41925031, 41630645, and 41521003) and the Startup Foundation for Introducing Talent of NUIST (1523142401062).

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Footnote

† Electronic supplementary information (ESI) available. See DOI: https://doi.org/10.1039/d4en00723a

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