Microplastics in river water: occurrence, weathering, and adsorption behaviour

Abstract

Microplastics (MPs) are emerging contaminants of concern due to their ubiquitous presence in aquatic environments and their ability to adsorb and transport other contaminants. In this study, the presence of MPs was determined in river water samples, reflecting their potential impact on the transport of other emerging contaminants in aqueous matrices. This study investigates the adsorption behavior of atrazine (ATZ), a widely used herbicide, onto pristine and UV-aged polyethylene (PE) and polypropylene (PP) MPs. The study revealed that UV aging enhances adsorption by increasing surface roughness and oxygen-containing functional groups. Batch adsorption experiments were conducted under varying environmental conditions, including pH, salinity, and dissolved organic matter changes. Adsorption kinetics were evaluated using pseudo-first-order, pseudo-second-order, and intraparticle diffusion models, with PSO providing a better fit, as indicated by lower p-values and higher R² values. The interparticle diffusion model showed that during the first stage of adsorption, surface adsorption was dominant, while pore diffusion was predominant at later stages. Desorption experiments indicated that aged MPs retain ATZ more effectively, reducing its potential for remobilization in aquatic systems. These findings provide insight into the environmental risks associated with MPs as carriers of pesticides and their implications for water quality and ecosystem health.

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Water impact

This study highlights the presence and critical role of microplastics (MPs) in transporting pesticides like atrazine in freshwater systems, emphasizing their impact on water quality and aquatic ecosystems. By demonstrating how MP aging enhances pollutant adsorption and reduces desorption, it provides valuable insights into the long-term environmental risks of MPs as carriers of toxic contaminants. These findings contribute to our understanding of contaminant behavior in aquatic systems, which is essential for assessing the ecological risks associated with MPs and their interactions with emerging contaminants.

1. Introduction

Microplastics (MPs), commonly defined as plastic fragments having a size smaller than 5 mm, have received considerable attention in recent times due to widespread distribution in various natural systems like groundwater,¹ oceans,² rivers,³ and sediments.⁴ Polyethene (PE) and polypropylene (PP) are the two most commonly found MPs in the environment.⁵ MPs may inevitably undergo one or more ageing processes (such as photooxidation or hydrolysis) during their lifecycle.⁶ The process of aging leads to chemical modifications of the surfaces of MPs, which affect characteristics such as oxygen-containing functional groups, specific surface area, and hydrophilicity.^7–9 These alterations may impact the capacity of these MPs to adsorb and desorb emerging contaminants in aqueous media. With the increase in the extent of MP pollution, evaluation of these characteristics has become very important to fully grasp the scope and severity of the issue. The weathering of MPs can also potentially affect the ability of MPs to transport emerging contaminants in aqueous media.¹⁰ Currently, the range of simulated MP ageing processes is limited; researchers generally rely on various accelerated ageing techniques including Fenton treatment,¹¹ ozonation,¹² freeze–thaw cycles,¹³ and high temperature oxidative ageing.¹⁴ As MPs move through environmental systems, they act as potential vectors of emerging contaminants like pesticides, pharmaceuticals, hormones, and other industrial chemicals.^15–17

Adsorption of emerging contaminants on MPs takes place through chemical and physical interactions such as electrostatic, hydrophobic, and other interactions.¹⁸ MPs accumulate pollutants and can transmit them to living organisms, which can lead to significant toxicological consequences.¹⁹ It is crucial to know the mechanism and process of contaminant adsorption onto MPs.²⁰ Emerging contaminants including pesticides are frequently detected in aquatic environments.²¹ Among the emerging contaminants frequently detected in water bodies, atrazine (ATZ) stands out due to its widespread use and persistence in aquatic systems. ATZ is extensively used as an herbicide for the effective and economical control of weeds in agricultural practices, owing to its low price and high efficiency. It has been prohibited in the European Union since 2003, but continues to be used in countries like India, China, the United States, and Brazil.¹⁰ ATZ is ubiquitous in aqueous media due to its persistent nature.²² ATZ exposure has been linked with adverse impacts on aquatic organisms and humans even at very low concentrations.²³ ATZ has the potential to interfere with the endocrine system, harm the cardiovascular, gastric, and reproductive systems, and cause cancer.^24,25 Thus, it is important to give special attention to ATZ due to its presence in aquatic environments, which enables it to interact with MPs.²⁶ However, there are very few studies on the interactions of ATZ and MPs and their ageing in the aquatic environment. Given the significant reliance on river systems for agriculture and drinking water in India, understanding MPs' role as carriers of contaminants like ATZ in these environments is crucial for water quality management. Thus, it is essential to look at the complex adsorption process of ATZ on MPs and understand the related processes. MP fibres were detected in the groundwater in Illinois, USA, with an average concentration of 15.2 particles per litre.²⁷ MPs have been identified in both surface and groundwater due to their ability to traverse through the soil pores and into groundwater reserves. Rivers have traditionally been seen as the primary channels by which MPs are transported to oceans. However, studies on the occurrences of MPs in rivers and the role of these in carrying other emerging contaminants is scarce in India. By focusing on ATZ, we not only address a compound of environmental relevance but also provide new insight into how MPs, particularly aged MPs, may influence the environmental transport and fate of persistent contaminants. While natural ageing of MPs is influenced by multiple factors such as mechanical abrasion, biodegradation, and chemical oxidation, UV-induced photooxidation has been widely used as a standardized method to simulate environmental weathering in laboratory due to its reproducibility and its ability to generate oxygen-containing functional groups on MP surfaces.

With these research gaps, the aim of this study was to identify the occurrences of MPs in river water. Further, this study fills the gap by systematically examining the adsorption behaviour between ATZ and MPs before and after ageing using adsorption isotherms and kinetics. The effect of UV ageing on the characteristics of PE and PP MPs was analyzed by FTIR, XPS, BET and zeta potential analysis. Additionally, the effect of solution chemistry like salinity, pH, and dissolved organic matter (DOM) on the adsorption was determined. The amount of ATZ desorbed from pristine and aged MPs in the aqueous matrix was determined and a possible mechanism for ATZ adsorption was proposed.

2. Materials and methods

2.1. Freshwater sampling locations and procedure

Grab samples of freshwater (3 L) were collected from a single location on the Kangsabati River (Gandhi Ghat Park), India, in December 2024 (Fig. S1). This river flows between Midnapore town (sampling location) and Kharagpur town, which is a highly and densely populated town (92 111.7 km⁻²), and it traverses both urban settlements and agricultural catchments, making it vulnerable to multiple sources of contamination. The sampling site was located near a wastewater treatment plant discharge point, which was deliberately chosen as a potential hotspot for microplastic contamination. While the results provide valuable insight into the characteristics of MPs in effluent-impacted zones, they are not intended to represent the entire river system. No independent control samples were included in this study. All experiments were conducted under identical conditions using the test samples, and the results were interpreted accordingly. This represents a limitation of the present work, as the absence of control experiments restricts the ability to fully isolate treatment-related effects.

Upon collection, the samples were digested with 25 mL, 30% H₂O₂ and 25 mL 0.05 M Fe(II)SO₄·7H₂O solution, and placed on a hot plate magnetic stirrer at 75 °C for 0.5 h and allowed to settle for 24 h.²⁸ Following digestion, the samples were passed through 0.2 μm glass fibre filters (Whatman) using a vacuum filtration unit. The samples collected on the filter were examined for particle sizes and shapes using bright-field microscopy. To confirm the identity of the particles as MPs, samples were dyed with Nile red and were observed under a fluorescence microscope. Stained particles were observed under orange red light with an excitation wavelength of 534–558 nm and emission wavelength of 590–646 nm. This analysis verified the MP particles and distinguished MPs from other debris or organic matter.

2.2. Materials

PE and PP virgin MPs were chosen as representative MPs owing to their widespread presence in aquatic environments. Virgin PE and PP pellets (Sigma Aldrich) were purchased, and grounded into powder using a mixer grinder. All grinding equipment was thoroughly cleaned with ethanol and rinsed with Milli-Q water before use to minimize the MP contamination. The powder particles were sieved with 75 to 600 μm sieves to obtain a uniform size range. ATZ powder was obtained from Sigma Aldrich, and the standard was prepared in ultrapure water. Acetonitrile, humic acid, hydrochloric acid (HCl), sodium chloride (NaCl), and sodium hydroxide (NaOH) were all of analytical grade and were purchased from Merck Life Sciences Pvt. Ltd, India.

2.3. MP ageing

The virgin PE and PP MPs were spread evenly on Petri dishes, and were subjected to photodegradation to simulate ageing due to UV exposure. A UV chamber was constructed using two 15 W UVA-340 lamps that can emit light from 400 to 300 nm to simulate sunlight. The MPs were exposed to 15 W UVA 340 lamps to simulate long UV rays. In order to mimic the sunlight cycle, the microplastics were exposed to UV rays for periods of 11 hours and kept in the dark for 13 hours. This wavelength range has a high level of intensity, making it more prone to photo-ageing compared to other wavelengths.⁷ The samples were positioned 15 cm from the light source, ensuring that the MP surface received the maximum light intensity of 1.8 mW cm⁻². The MP particles were agitated every 24 hours to maintain uniform aging.

2.4. MP characterization

The surface morphology of both virgin and aged MPs was analyzed using Scanning Electron Microscopy (FESEM, Zeiss Merlin, Germany).²⁰ Prior to SEM analysis, the samples were washed using Milli-Q water and then oven dried at 60 °C for a duration of 24 hours. The surface elements of the MPs were analysed using X-ray photoelectron spectroscopy (XPS, ULVAC PHI (Physical Electronics), USA).²⁹ To observe the influence of surface charges of MPs and atrazine solution on the adsorption mechanism, zeta potential was measured. The zeta potential of atrazine and MPs was determined in an aqueous medium using a Zetasizer Nano ZS90 (Malvern, UK) instrument.²⁹ The X-ray diffraction (XRD) study was performed to understand the crystallinity of MPs using an XRD diffractometer (D2 Phaser, Bruker, USA). The sample scanning range spanned from 10 to 80°, with a scanning rate of 5° min⁻¹.²⁹ The data were analyzed using HighScore Plus software. The functional groups of the virgin and aged MPs were determined using FTIR (Alpha II, Bruker, Massachusetts, USA) spectra with the spectral range of 4000–500 cm⁻¹.²⁰ The pore volumes and specific surface areas of the MP samples were measured with a Surface Area & Porosimetry System (ASAP2460; Micromeritics, USA).³⁰ Prior to the analysis, MP samples were degassed for 8 h at 80 °C. Further, N₂ adsorption and desorption tests were conducted at −196 °C.

2.5. Determination of atrazine

Samples collected at different intervals of adsorption were analyzed using an Ultra High-Performance Liquid Chromatograph (Shimadzu, Kyoto, Japan) with a PDA detector. The experimental design consisted of three replicates and the obtained data were reported as the mean ± standard deviation. The detection technique was based on a sample injection volume of 20 μL, and the wavelength was set at 220 nm. The mobile phase was acetonitrile and water in a 70 : 30 (V/V) ratio, with a flow rate of 1.0 mL min⁻¹.

2.6. Batch adsorption experiments

Batch adsorption studies were performed to evaluate whether MPs can act as carriers of ATZ in aqueous media. The studies were conducted in 30 ml flat bottom amber glass vials sealed with threaded caps. PE and PP MP samples were placed into the vials followed by addition of ATZ solutions at various concentrations (1, 3, 5, 10, 15, and 20 mg L⁻¹). ATZ solutions were prepared using Milli-Q water to evaluate the intrinsic adsorption behavior of the adsorbent in the absence of competing species. The mixture was incubated at a specific temperature of 25 °C for 48 hours with continuous oscillation at 200 rpm. Samples were taken out at various time intervals (0.5, 1, 3, 5, 8, 12, 24 and 48 h) and analyzed for ATZ concentration. The initial pH of the solution was adjusted to 11, 9, 7, 5, and 3 using 0.1 M HCl and NaOH solution. Although the pH of natural river water typically near neutral, a broader pH range was tested to evaluate the influence on surface charge and the impact of pH on adsorption behaviour. The impact of salinity on the adsorption was also investigated to simulate estuary and marine environments. Sodium chloride (NaCl) was used to alter the salinity of the solution, with concentration ranging from 5–35 g L⁻¹. In natural water sources, DOM is often present and can affect the adsorption behaviour of contaminants. The impact of DOM on the adsorption capacity was also assessed to ensure that the adsorption process is studied under more realistic conditions. Humic acid (HA) was added to replicate DOM at a concentration between 5 and 20 mg L⁻¹. The adsorption capacity was calculated using the following formula:where q_t (mg g⁻¹) is the amount adsorbed at time t; C₀ (mg L⁻¹) and C_e (mg L⁻¹) represent the initial and equilibrium concentrations of ATZ, respectively; V (mL) is the solution volume; and m (mg) is the adsorbent mass.

2.7. Adsorption models

The ATZ adsorption capacity of both MPs was determined by Langmuir, Freundlich, and Henry isotherm equations. The study used the pseudo-first-order (PFO) kinetic model, pseudo-second-order (PSO) kinetic model and intraparticle diffusion model to describe the adsorption kinetics of ATZ on MPs, and the equations used for model-fitting are:

Henry isotherm model:

q_e = k_HC_e (2)

Langmuir isotherm model:Freundlich isotherm model:PFO model:

ln(q_e − q_t) = ln q_e − k₁t (5)

PSO model:Intraparticle diffusion model:

q_t = k_pt^1/2 + C (7)

Here, q_e represents the equilibrium adsorption capacity (mg ATZ per g of MP); k_d is the partition constant (L g⁻¹); k_L is the Langmuir constant (L mg⁻¹); q_m represents the maximum capacity of adsorption, the constants K_f and 1/n are related to the Freundlich isotherm, k₁ and k₂ are the first (1 min⁻¹) and second (g mg⁻¹ min⁻¹) order rate constants, respectively; C is the constant related to the boundary layer thickness, and k_p (g mg⁻¹ min^0.5) denotes the rate constant of the intraparticle diffusion model. The R² and p-values were used to determine the best-fit for the abovementioned data. The statistical analysis of experimental data was conducted using Microsoft Excel and graphed using Origin 2024 software.

2.8. Desorption assessment

The ATZ desorption study was conducted after achieving equilibrium in adsorption experiments. The MPs were filtered after the adsorption reached equilibrium conditions. The collected MPs were dried and mixed with Milli-Q water in glass tubes. The tubes were incubated for 48 h (equilibrium time) at 25 °C in a shaker at 200 rpm. The Milli-Q water was collected after 48 h, filtered using 0.2 μm glass fiber syringe filters, and injected in the HPLC system for ATZ concentration determination.

2.9. Quality assurance and quality control (QA/QC)

QA/QC protocols were implemented at every stage, to mitigate potential contamination. Water samples were stored in glass jars, which were pre-washed and rinsed with Milli-Q water to eliminate any potential contamination. During sample collection, strict precautionary measures were taken to prevent contamination, such as minimizing exposure to air, synthetic fabrics, and plastic materials. The atrazine calibration curves were prepared using standard ATZ solutions with correlation coefficients (R²) greater than 0.999. Method blanks and procedural blanks were analyzed to ensure the absence of contamination. All the samples were measured in triplicate to ensure reproducibility.

3. Results and discussion

3.1. Presence of microplastics in river water

MPs were found in all three water samples collected from the river averaging 8 particles per L. Some similar studies have been conducted in other Asian countries like Nepal, Maldives and Sri Lanka, to examine the occurrence of MPs in freshwater sources.³¹ In Nepal, the MP concentration found in freshwater resources ranged from 2.96–65 particles per L.³² A quantitative investigation identified the occurrences of MPs in sediments and surface water of the Mahananda River in the range of 167–193 particles per g and 59–100 particles per l, respectively.³³ In another study, MPs in a freshwater lake named Vellayani Lake, situated in Kerala, India, were identified in the range of 20 to 100 particles per m³, with an average of 65 particles per m³.³⁴ However, during summer, the concentration reduced to 34 to 67 particles per m³, with an average of 49 particles per m³. The authors concluded that the abundance of PE and PP MPs is due to the disintegration of single-use disposable plastics.³⁴ A study conducted in the presence of MPs in the Valvanti River, India, revealed the occurrence of MPs in the range of 1.1 to 7.5 MPs L⁻¹, with an average (± SD) of 2.91 ± 1.69 MPs L⁻¹.³⁵ Most MPs found were fibres (96.9%) and colourless (70.49%). The risk assessment study revealed that despite being moderately impacted by MPs, the river is subject to high ecological risk.³⁵ A review of MP pollution in Indian freshwater systems, including 198 studies from 1990–2025, showed that MPs ranged from 5.9 particles per L to 846 ± 136 particles per L.³⁶ The primary sources of these MPs were found to be plastic mulching in agriculture, industrial discharge, urban waste streams, and untreated sewage.³⁶ MP concentration in Sabarmati River sediments was determined, and it was seen that the MP size ranged from 75 μm to 4 mm.³⁷ The authors found that the MPs were absent upstream of Ahmedabad city and progressively increased when approaching the city, with the highest concentration reported near the Pirana waste dumping site.³⁷

This suggests that our study area is moderately impacted by microplastic pollution. MPs were found in different shapes (among which fragments and fibres were the most abundant). The presence of MPs creates concerns about their possible ecological impacts, including ingestion by aquatic organisms and their role as vectors of adsorbed contaminants. Localized plastic pollution and wastewater treatment plant effluent discharge in the river, along with fishing nets, were found to be the probable causes of MP pollution. A map of the sampling locations is provided, showing the wastewater treatment plant discharge point, adjacent agricultural fields, and major urban runoff channels contributing to the river (Fig. S1). These plastics undergo physical, chemical, and biological degradation processes, breaking down into smaller fragments. The continuous input from these sources, combined with the river's flow, can spread MPs downstream, impacting not only local aquatic ecosystems but also larger connected water bodies like estuaries and oceans.

3.2. Alteration in microplastic characteristics

The physicochemical parameters (i.e., pore size, surface area, structure, elemental composition, and hydrophobicity) of MPs may be substantially modified during ageing processes. The alterations observed before and after UV ageing of MPs are discussed below.

3.2.1. Surface morphology. The SEM images of virgin and aged PE and PP MPs are shown in Fig. 1. Both pristine MPs exhibit relatively non-porous, smooth surfaces. In contrast, aged PE MPs display new wavy, irregular folds, layered structures, and cracks. Similarly, the UV-aged PP MPs show the formation of new cracks and a wrinkled surface, indicating fragmentation of the MPs.

Fig. 1 SEM images of the microplastics studied.

The pore volume and surface area of MPs affect the adsorption process by determining the sites available for adsorption. MPs with a greater specific surface area provide more active adsorption sites and enhance the capacity of MPs to adsorb contaminants. As presented in Table 1, pristine and aged PE and PP MPs exhibited different surface properties. The surface areas of both aged MPs were greater than that of pristine MPs, which is similar to that observed in SEM images. As the PE MPs had a greater specific surface area than PP MPs, they showed higher adsorption capacity. These changes imply that aged MPs may pose a greater environmental risk compared to pristine MPs due to their enhanced potential to adsorb and transport contaminants.

Table 1 Specific surface area and pore volume of the MPs studied

Polymer type	Size (μm)	Specific surface area (m^2 g^−1)	Pore volume (cm^3 g^−1)
Pristine PE MPs	<150	2.576	0.004
Aged PE MPs	<150	2.634	0.004
Pristine PP MPs	<150	0.289	Not detected
Aged PP MPs	<150	0.459	0.001

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3.2.2. FTIR spectroscopy. The alterations in the FTIR spectra of MPs after UV ageing and adsorption of ATZ are shown in Fig. 2. Characteristic peaks of PE MPs were observed in virgin PE MP samples, i.e., stretching vibration of CH₂ and CH₃ groups at 2800–2900 cm⁻¹, and rocking and bending vibrations of CH₂ groups were observed at 720 and 1460 cm⁻¹, respectively (Fig. 2).

Fig. 2 FTIR spectra of (a) PE MPs and (b) PP MPs.

The bands found at 3112, 1614, and 718 cm⁻¹ in MP samples after adsorption likely result from the stretching vibrations of C–H groups, N–H bending, and C–Cl stretching of ATZ, respectively. Prominent changes were not observed for the characteristic peaks of the photodegraded samples. However, subtle changes in the peak shapes and absorbance intensities were noted. The formation of the carbonyl peak (1715 cm⁻¹) while observing the early changes during the photodegradation of PE and PP has been reported by several researchers.^38–40 As this study was conducted for approximately 30 days, only the early changes were visible in the FTIR spectrum. The characteristics peaks of PP MPs were observed in virgin and aged MPs (Fig. 2(b)). In the case of aged MPs, peaks in the hydroxyl region (1700 cm⁻¹) and minor shifts in CH₂ bending bands (1450–1460 cm⁻¹) confirm PP oxidation. Peaks at 1614 cm⁻¹ and changes in the 3500–3300 cm⁻¹ region confirm adsorption of atrazine onto PP.

3.2.3. X-ray photoelectron spectroscopy. X-ray photoelectron spectroscopy (XPS) furnishes strong evidence of degradation in both PP and PE samples with significant change in their carbon and oxygen atomic percentage. PP MP samples showed significant chemical changes after the degradation process. The carbon content reduced from 98.92% to 96.69% indicating a decrease of 2.23% (Fig. 3(c) and (d) inset). Conversely, the oxygen concentration increased from 1.08% to 3.31% (Fig. 3(c) and (d)).

Fig. 3 XPS spectra of PE MPs: (a) pristine MPs and (b) aged MPs; PP MPs: (c) pristine MPs and (d) aged MPs.

The XPS scan of virgin PE and PP samples showed two primary peaks correlating to C 1s and O 1s. The virgin PE and PP samples exhibit a hydrocarbon surface with high carbon content, while the subsequently degraded samples display a significant increase in oxygen content. Both PP and PE samples show significant chemical changes after the degradation process. The carbon content in PE samples decreased from 99.72% to 98.96% after degradation, while the oxygen content increased from 0.28% to 1.04% (Fig. 3(a) and (b)).

The XPS spectrum of the C 1s peak provided a detailed insight into the photo-oxidation process. The virgin PE and PP samples have a single distinct peak at approximately 284 eV, indicating the presence of C–C or C–H bonds in the polymer backbone. Deconvolution of the peaks obtained from the degraded samples yielded three constituent peaks. One main peak at 284 eV, representing the C–C, C–H bonds, another shoulder peak at 286 eV indicating C–O bonds, and a smaller peak at 288 eV suggestive of C=O bonds. The convoluted peaks, therefore, substantiate the oxidative process during the photodegradation. It is evident that the primary peak at 284 eV (C–C/C–H) is retained as photodegradation does not impact the polymer backbone. However, the development of peaks at 286 and 288 eV suggests the formation of C–O and C=O groups, respectively.

It can be inferred from the results that both polymers undergo significant surface oxidation. PP MP samples showed higher susceptibility to oxidation, evidenced by higher oxygen incorporation, more pronounced development of the oxidised carbon species, and greater reduction in the primary carbon content. These findings corroborate FTIR spectral data regarding the oxidation of the carbon backbone (1715 cm⁻¹). Complementary surface specific information can be observed from the XPS data confirming the oxidation of the polymer surface during UV radiation. Higher susceptibility to oxidation was observed in PP than PE; this might be due to the tertiary carbon atoms present in the backbone of the polymer surface.⁴¹

3.2.4. X-ray diffraction. Fig. 4 shows the characteristic peaks of the PE and PP MPs. The sharp diffraction peaks observed in both MPs indicate their inherent crystallinity, which significantly influences the adsorption performance of ATZ onto these MPs. The peaks observed at 21.43° and 23.72° were the primary crystalline peaks identified for PE MPs. PP MPs exhibited distinct peaks at 2θ values of 14.13°, 16.94°, 18.65°, 21.27°, and 25.69°. Aging reduced the crystallinity of both materials, as evidenced by subtle changes in peak intensity and width. Specifically, the crystallinity decreased by 10% in PE MPs and 7% in PP MPs. This reduction can be attributed to UV irradiation, which disrupts polyethylene chains and decreases the molecular weight, thereby diminishing the crystalline regions.

Fig. 4 XRD spectra of (a) pristine and aged PE and (b) pristine and aged PP microplastics (inset: magnified view of the primary crystalline peaks).

3.3. Microplastics as carriers of atrazine

3.3.1. Effect of adsorption time. The adsorption process is not instantaneous. Over time, the adsorbate (e.g., a contaminant like ATZ) gradually attaches to available adsorption sites on the adsorbent (e.g., MPs). Adequate contact time allows the adsorbate to diffuse and interact with the adsorbent, eventually reaching equilibrium. The adsorption kinetics, or the rate at which adsorption occurs, depends on the contact time. In the early stages, adsorption typically occurs rapidly as surface sites are readily available. Over time, the adsorption rate slows as these sites become occupied, requiring more time for the adsorbate to find available or deeper sites within the adsorbent. The impact of contact time on ATZ adsorption onto different MPs was investigated. The contact time was established to compare the outcomes of several studies, since it is essential to obtain a consistent equilibrium time applicable to all experiments (Fig. 6(a)). The adsorption rate increased significantly during the first 12 hours, but little change in ATZ adsorption (±0.0013 mg g⁻¹) was observed after 24 hours. Therefore, the remaining adsorption studies were conducted over a 24-hour period. In this study, adsorption equilibrium times were experimentally determined for pristine MPs and used for both pristine and aged MPs. This method facilitates comparative evaluation but assumes that aging does not substantially affect equilibrium time.

3.3.2. Effect of initial atrazine concentration. The initial ATZ concentration was altered to assess its impact on adsorption by MPs. The quantity of ATZ adsorbed on MPs increased with an increase in the initial ATZ concentration (1–15 mg L⁻¹) (Fig. 6(b)). For all types of MPs studied, the adsorption capacity increased sharply as the ATZ concentration rises from 1 to 5 mg L⁻¹. This is because, at higher concentrations, more free ATZ molecules are available to interact with available adsorption sites on the MPs' surfaces. After reaching a concentration of approximately 10 mg L⁻¹, the adsorption capacity starts to plateau or decrease slightly. This trend indicates that the MPs' surfaces are nearing saturation, with fewer available sites remaining for additional ATZ molecules to adsorb. Between the two MPs studied, PE, especially aged PE, emerges as a more probable carrier for ATZ compared to PP, potentially due to its chemical composition and surface properties. PE, being a linear hydrocarbon polymer, undergoes more uniform photo-oxidation along its backbone during ageing. In contrast, PP MPs contain bulky methyl side groups, which sterically hinder oxidation and reduce the extent of functionalization during UV aging. The results indicate that ATZ adsorption was highest at a concentration of 10 mg L⁻¹, where the MPs can still offer enough free adsorption sites.

3.3.3. Effect of microplastics dose. The impact of MP dosage on ATZ adsorption was evaluated and is presented in Fig. 6(c). The dosage of MPs was altered from 1 to 4 g L⁻¹, maintaining the initial ATZ concentration constant at 10 mg L⁻¹. The ATZ adsorbed on MPs increased with an increase in MP dosage due to the larger surface area available for adsorption.¹⁸ The adsorption capacity increased sharply between 1 and 2.5 g L⁻¹ of MP dosage, which plateaued between 2.5 and 4 g L⁻¹. This suggests that while increasing the dose raises the total surface area, adsorption slows once a critical dose is reached, likely due to fewer free ATZ molecules or saturation of available MP surface sites.⁴² Although MPs exhibit a high negative surface charge that should favor electrostatic repulsion, aggregation may still occur due to environmental factors. Increased ionic strength can compress the double layer and reduce electrostatic repulsion, enabling van der Waals or hydrophobic interactions to dominate thereby reducing the available surface area for adsorption.^43,44

Aging of both PE and PP significantly enhanced their adsorption capacities with aged PE showing the most significant improvement. This enhancement is possibly due to the aging process that introduces surface oxidation or functional groups, increasing available adsorption sites and improving the affinity between ATZ and the MPs. In addition, aging can lead to changes in surface roughness and pore size as confirmed by SEM images and BET study. This structural alteration increases the number of accessible adsorption sites and facilitates better diffusion of ATZ molecules into the MPs. Together, these physical changes significantly improve the adsorption efficiency of aged MPs compared to pristine MPs. Based on these findings, an optimal MP dose of 2.5 g L⁻¹ was chosen for further experiments.

3.3.4. Effect of pH. The influence of initial solution pH (3–11) on the adsorption was examined and is depicted in Fig. 6(d). The findings indicate that solution pH had specific impact on ATZ adsorption by the MPs. As depicted in SEM images (Fig. 1) and surface area analysis (Table 1), aged MPs developed a flaky, rough morphology and increased specific surface area compared to pristine MPs. Concurrently, zeta potential measurements indicated a lower surface charge for aged MPs. With the increase in solution pH from 7 to 11, the adsorption capacity showed a slight downward trend. Solution pH did not have any major impact on adsorption density when the pH of the solution was altered from 3 to 7. Notably, aged PP MPs exhibited lower sensitivity to pH values. At pH 3, the surface charge of ATZ is lowest leading to less electrostatic repulsion between ATZ and the MPs as shown in Fig. 5(a) and (b). Since the MPs also have weak surface charges at pH 3, there is less interference in the adsorption process compared to higher pH values where more pronounced repulsion or competition from other ions may occur. ATZ has a weakly polar nature and is not significantly ionized over a broad pH range (pH 4–9).⁴⁵

Fig. 5 Zeta potential of (a) MP particles and (b) ATZ.

This means that its chemical structure and hydrophobicity do not change drastically across typical environmental pH values, leading to relatively stable adsorption behaviour regardless of pH shifts. This research established that the adsorption of ATZ to MPs remained relatively constant throughout a pH range of 3 to 11, suggesting that ATZ adsorbed in MPs may present similar risks to aquatic organisms in natural environments with varying pH values.

3.3.5. Effect of salinity. As MPs move from rivers to oceans, the salinity of the aqueous matrix progressively alters with it. The impact of salinity on atrazine adsorption to MPs was assessed, since MPs ultimately enter the ocean via drainage systems and rivers. As shown in Fig. 6(e), salinity was varied from 0.5–3.5% to simulate fresh water to ocean water salinity. One set of controls using 0% salinity (Milli-Q water) was used to compare results. Fig. 6(e) illustrates that an increase in salinity had no substantial impact or a slight decrease in the adsorption capacity. The highest adsorption rate was seen at a NaCl concentration of 0.5%, whereas the adsorption capacity remained constant at NaCl concentrations of 2.5 and 3.5%. The increase in ionic strength of the solution increases the concentration of Na⁺ ions in the aqueous solution, which adsorbs onto the surface of negatively charged MPs, competing for the available adsorption sites with the pesticide molecules.

Fig. 6 Adsorption studies with ATZ and different MPs: effect of (a) adsorption time, (b) initial atrazine concentration, (c) microplastic dose, (d) initial pH, (e) salinity, and (f) humic acid.

It was observed that the adsorption capacity for four pesticides namely malathion, difenoconazole, carbendazim, and diflubenzuron increased with the increase of NaCl concentration initially and remained constant at higher concentration.⁴⁶ Nonetheless, salinity does not demonstrate a consistent pattern in its influence on adsorption.⁴⁷

3.3.6. Effect of dissolved organic matter. NOM is widely present in natural water matrices and contains various functional groups including hydroxyl, carboxyl, and phenolic groups. In this study, HA was used as a representative of DOM, and its concentration was varied from 0 to 25 mg L⁻¹ (Fig. 6(f)). As the HA concentration increased, the adsorption capacity of ATZ onto pristine PE MPs significantly decreased. Specifically, the adsorption capacity of pristine PE MPs dropped from 0.37 ± 0.023 mg g⁻¹ to 0.17 ± 0.168 mg g⁻¹ as the HA concentration increased from 0 to 25 mg L⁻¹. For aged MPs, a similar decreasing trend was observed with increasing HA concentration. Specifically, the adsorption capacity of aged PE MPs decreased from 0.42 ± 0.031 mg g⁻¹ at 0 mg L⁻¹ HA to 0.41 ± 0.027 mg g⁻¹ at 25 mg L⁻¹ HA, while aged PP MPs dropped from 0.32 ± 0.121 mg g⁻¹ to 0.36 ± 0.019 mg g⁻¹ over the same HA range.

Our previous study reported DOM levels of ∼5 mg L⁻¹ in the same river location.⁴⁸ This reduction in adsorption capacity can be due to several factors. As HA concentration rises, more adsorption sites on the MPs become occupied by HA, leaving fewer available for ATZ. Additionally, HA can form complexes with ATZ in solution, altering the physical and chemical properties of ATZ and reducing its affinity for adsorption onto MPs. These complexes increase the solubility of ATZ in water, weakening its hydrophobic interaction with the MPs. Furthermore, HA, which is negatively charged, can enhance electrostatic repulsion between the MPs and ATZ, further reducing ATZ adsorption. In some instances, DOM may associate with the adsorbent (MPs) prior to interacting with other emerging contaminants. Similar results have been reported previously where increased NOM concentrations led to reduced pesticide adsorption capacity due to competition for the same adsorption sites.^46,49

3.4. Adsorption isotherms

The adsorption characteristics were further analysed by fitting the adsorption data to the Langmuir, Freundlich, and Henry isotherm models. The isotherm models for the sorption of ATZ onto different MPs are depicted in Fig. 7 and the associated fitting parameters are listed in Table 2. The results showed that the Langmuir isotherm model effectively described the ATZ adsorption onto different MPs with higher R² values than the other two models. The Langmuir affinity constant (K_L) was higher for PE MPs in comparison to PP MPs. Similarly, in both cases UV aged MPs exhibited elevated K_L values compared to pristine MPs. The higher K_L values observed for aged MPs indicate a stronger binding affinity towards ATZ molecules. This is consistent with the physicochemical changes observed in aged MPs. Both the aged MPs showed lower R² values for the Langmuir isotherm indicating that only the Langmuir model may not be sufficient to fully describe the adsorption process.⁵⁰

Fig. 7 Isotherm models of ATZ on different MPs: (a) Langmuir, (b) Freundlich, and (c) Henry model.

Table 2 Parameters derived from different isotherm models for the adsorption of ATZ onto different MPs

		Pristine PE	Aged PE	Pristine PP	Aged PP
Langmuir	q m	0.6143	0.6686	1.2962	1.1395
	K L	0.3265	0.5413	0.0292	0.0619
	R ^2	0.94	0.86	0.98	0.91
Freundlich	K F	0.1828	0.2849	0.0449	0.1039
	n	2.7949	4.1649	1.3323	2.0296
	R ^2	0.72	0.43	0.93	0.63
Henry	K H	0.0308	0.0351	0.0222	0.0238
	R ^2	0.75	0.62	0.96	0.75

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The lower R² values observed in the Henry model may be due to the complex interactions at higher concentrations of ATZ. The results of the adsorption study imply that surface ageing affects adsorption capacities, although both the studied MPs favoured simpler, monolayer adsorption behaviour. The low R² values obtained for the Henry model indicate that the ATZ adsorption onto MPs does not follow a simple linear relationship, suggesting deviations from ideal partitioning behaviour. However, aged MPs exhibited a steeper slope, indicating enhanced adsorption due to surface oxidation and functional group formation.

3.5. Adsorption kinetics

The experimental data of ATZ adsorption onto different MPs were used to fit in PFO, PSO and IPD models and are shown in Fig. 8. The relevant adsorption kinetic parameters of ATZ adsorption on different MPs were calculated and listed in Table 3. High R² values were observed for both PFO and PSO models, while a slightly higher R² value was observed for the PSO model. This result indicates that chemisorption is possibly the dominant mechanism behind this adsorption process.⁵¹ Additionally, the higher R² values observed in the PFO model suggest that the adsorption is also associated with physisorption.⁵² Furthermore, p-values obtained from linear regression analysis were low for both models, indicating statistically significant fits.

Fig. 8 Adsorption kinetics of ATZ: (a) pseudo-first-order, (b) pseudo-second-order and (c) intra-particle diffusion model.

Table 3 Kinetic characteristics of ATZ adsorption on different MPs

MPs	Pseudo-first order model				Pseudo-second order model
	K 1	q e	R ^2	P-Value	K 2	q e	R ^2	P-Value
Pristine PE	0.1767	0.3297	0.96	1.46 × 10^−05	0.7457	0.4585	0.99	7.84 × 10^−06
Aged PE	0.1769	0.4237	0.97	1.03 × 10^−05	0.6375	0.5409	0.98	1.52 × 10^−06
Pristine PP	0.1655	0.1693	0.99	1.1 × 10^−07	1.7601	0.2352	0.99	6.46 × 10^−07
Aged PP	0.1732	0.2493	0.99	4.14 × 10^−08	1.2182	0.3448	0.99	3.46 × 10^−08

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The IPD model revealed that ATZ adsorption involves three distinct phases, as shown by the three different stages in Fig. 8(c). In the first stage, a steeper gradient was observed in comparison to the second phase, irrespective of the type of MPs or their aging and this first phase is the bulk diffusion phase. The steeper slope in the first phase indicates a rapid transfer of ATZ from the bulk solution to the external MP surface. The decrease in slope in the second phase denotes ATZ diffusion into the MP pores, as the accessible adsorption sites on the surface were filled in the first stage; this is pore diffusion.²⁰ The lower R² values and reduced slope gradient in the third stage (adsorption) imply that IPD is no longer the rate-limiting step, and adsorption has reached equilibrium, with the adsorption rate slowing down.⁵³ In addition, aging of MPs has been shown to enhance adsorption capacity in the early stages.

3.6. Desorption assessment

The transport and toxicity of emerging contaminants are inherently connected to the desorption characteristics of MPs in aqueous media. Desorption experiments, conducted in a controlled environment as described in section 2.7, were used to evaluate the potential of MPs to carry ATZ in water bodies. Among all the MPs studied under identical conditions, pristine PE particles showed the highest ATZ release percentage followed by aged PE. The desorption percentage was observed to increase over time, reaching its highest value at 48 hours (Fig. 9). Both aged PE and PP MPs exhibited lower desorption rates compared to the pristine MPs. The aging of MPs likely introduces new binding sites and increases surface roughness, enhancing adsorption affinity and making ATZ desorption more challenging.⁵⁴ This phenomenon aligns with the observation that a higher adsorption capacity of aged MPs generally translates to a slower desorption rate, as ATZ molecules are more strongly held on the surface of aged MPs.

Fig. 9 Desorption rate of ATZ from different MPs.

These results indicate that aquatic organisms may be at increased risk due to ingestion of ATZ-loaded MPs. Overall, MPs can act as transporters of ATZ into aquatic ecosystems and biological tissues through the adsorption–desorption process.

3.7. Adsorption mechanism

The possible interaction mechanisms of ATZ and MPs were proposed based on the characterization of MPs and environmental factors. As ATZ molecules have negative surface charge across the pH range of 3 to 11 (Fig. 5), the adsorption of ATZ on MPs is likely to be influenced by electrostatic interactions, even though ATZ is primarily hydrophobic. The negative charge of ATZ increases with an increase in solution pH, which decreases the adsorption capacity (Fig. 5(b)). This phenomenon observed might be due to the increased electrostatic repulsion between negatively charged MPs and increased negative charge of ATZ at higher pH.

Since ATZ is hydrophobic in nature, hydrophobic interactions are most likely to be the key mechanism behind the adsorption onto both pristine and aged MPs. Hydrophobic pollutants like ATZ tend to attach to hydrophobic surfaces of MPs, thereby minimizing their interactions with aqueous matrices.³⁰ In the presence of NOM, the surface properties of the MPs remain unchanged. However, NOM molecules may compete with ATZ for adsorption sites by forming hydrogen bonds or occupying free surface sites on MPs. MP ageing also significantly impacts the adsorption process by increasing surface roughness and pore size in both PE and PP MPs. The rougher, more porous surface of aged MPs provides additional sites for physical adsorption of ATZ enhancing adsorption capacity. Furthermore, the oxidation of the microplastic surface and formation of carbonyl/carboxyl groups can provide potential spots for hydrogen bonding, allowing weak interactions between ATZ and MPs through these functional groups. The XRD spectrum showed reduction in crystallinity of aged MPs; this confirms a more active surface area in aged MPs which increased the adsorption capacity of aged MP particles. The aged MPs present a higher negatively charged surface due to the increased oxygen content as seen from XPS analysis, potentially increasing the adsorption capacity of aged MPs.

4. Conclusions

This work highlighted the critical role of both aged and pristine MPs in transporting pesticides within aqueous matrices under ambient environmental conditions. Based on the significant findings of this research, the following key conclusions have been derived:

1) Fragments and fibres were the most commonly found shapes of MPs in river water samples. These MPs are contaminants of concern due to possible ecological effects, and their role as vectors for adsorbed contaminants.

2) The aging of MPs altered the surface morphology and area of MPs, producing various oxygen-containing functional groups. These changes increase the potential of MPs to carry pesticides in aqueous environments.

3) The isotherm model fitting showed that the adsorption process of ATZ onto MPs involved both chemisorption and physisorption. The kinetic studies indicated that the adsorption of ATZ followed PSO kinetics and a strong correlation was observed between predicted and observed values by regression analysis.

4) Zeta potential and other analysis revealed that electrostatic interactions were not the only responsible factor for adsorption, indicating the role of various other factors like surface adsorption, pore filling, hydrophobic interactions, and hydrogen bonds.

5) This study underscores the importance of ATZ–MP interactions in shaping their environmental transport dynamics. Additionally, it offers crucial understanding into the behavior and mechanism of ATZ adsorption, which can help in preparing effective strategies for mitigating and managing MPs and other emerging contaminants.

Conflicts of interest

The authors declare that they have no known competing financial interests or personal relationships that could have appeared to influence the work reported in this paper.

Data availability

All data supporting the findings of this study are available within the paper and its supplementary information (SI).

Supplementary information is available. See DOI: https://doi.org/10.1039/d5ew00614g.

Acknowledgements

The authors wish to thank all technical staff of Central Research Facility (CRF), Indian Institute of Technology Kharagpur, India for their technical support during XPS, XRD, SEM analysis. The authors acknowledge Rajesh Kola for helping with FTIR analysis.

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