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Article

Antibiotic Adsorption by Microplastics: Effect of Weathering, Polymer Type, Size, and Shape

by
Thomas Easton
1,
Vaibhav Budhiraja
2,
Yuanzhe He
1,†,
Qi Zhang
1,†,
Ayushi Arora
3,
Vasileios Koutsos
4 and
Efthalia Chatzisymeon
1,*
1
School of Engineering, Institute for Infrastructure and Environment, The University of Edinburgh, Thomas Bayes Road, Edinburgh EH9 3FG, UK
2
Department of Polymer Chemistry and Technology, National Institute of Chemistry (Kemijski Inštitut), Hajdrihova 19, 1000 Ljubljana, Slovenia
3
School of Chemistry, Joseph Black Building, University of Edinburgh, David Brewster Road, Edinburgh EH9 3FJ, UK
4
School of Engineering, Institute for Materials and Processes, The University of Edinburgh, King’s Buildings, Robert Stevenson Road, Edinburgh EH9 3FB, UK
*
Author to whom correspondence should be addressed.
†
These authors contributed equally to this work.
Environments 2025, 12(4), 120; https://doi.org/10.3390/environments12040120
Submission received: 25 February 2025 / Revised: 7 April 2025 / Accepted: 8 April 2025 / Published: 12 April 2025
(This article belongs to the Special Issue Advanced Research on Micropollutants in Water, 2nd Edition)
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Abstract

:
The interaction of microplastics (MPs) with organic micropollutants, such as antibiotics, facilitates their transport in aquatic environments, increasing mobility and toxicological risk. The diverse polymer types, sizes, and shapes in wastewater present a challenge in understanding the fate of persistent organic micropollutants. This study examines ceftazidime adsorption on five polymer types—polyethylene terephthalate (PET), polyethylene (PE), hard and soft polystyrene (PS), hard and soft polyurethane (PU), and tyre wear particles (TWPs, including three passenger tyres and one truck tyre) in various forms (fibres, beads, foam, and fragments) and sizes (10–1000 µm). MPs underwent weathering (alkaline hydrolysis, UVC-activated H2O2, and Xenon lamp irradiation) to simulate environmental conditions. Their physical and chemical changes were analysed through mass loss, carbonyl index, scanning electron microscopy, and atomic force microscopy. The adsorption values (mg g−1) for pristine and weathered MPs, respectively, were as follows: PET (0.664 and 1.432), PE (0.210 and 0.234), hard PS (0.17 and 0.24), soft PS (0.53 and 0.48), hard PU (0.19), soft PU (0.17), and passenger TWPs—Bridgestone (0.212), Michelin (0.273), Goodyear (0.288), and Kumho truck TWPs (0.495). The highest and lowest adsorption were observed in weathered PET (1.432 mg g−1) and pristine hard PS/soft PU (0.17 mg g−1), respectively. Sorption kinetics and isothermal models showed that aged MPs exhibited higher sorption due to surface cracks, fragmentation, and increased adsorption sites. These findings enhance scientific knowledge of MP–antibiotic interactions in wastewater and can underpin studies to mitigate MP pollution and their adverse effects on the environment and humans.

1. Introduction

Microplastics (MPs) are present in wastewater treatment systems where they accumulate persistent organic pollutants, therefore amplifying their environmental and health risks due to increased ecotoxicity and endocrine disruption of aquatic ecosystems [1]. When organic pollutants, such as antibiotics, are adsorbed into MPs, their adverse environmental and health effects are exacerbated. The harmful effects of the excessive use of antibiotics and the development of antibiotic resistance in global water resources and in wastewater treatment plants (WWTPs) have been demonstrated over the last decades [2]. The accumulation of antibiotics in MPs can therefore substantially increase risks on public health and the environment. Interactions between antibiotics and MPs are important to consider in the framework of an effective risk evaluation of MPs. The large variety of MP types with different sizes and shapes in wastewater presents a complex challenge for understanding such interactions. Different studies have indicated that there is no clear pattern between the size distribution and shape of MPs found in WWTPs. For example, Murphy et al. reported that there are different shapes of MPs present in wastewater that include fragments (67.3%), fibres (18.5%), film (9.9%), beads (3.0%), and foam (1.3%) [3]. In addition to this, the composition varies throughout the stages in WWTPs, with significant variation across sampling sites.
A prominent type of MPs is fibres, which are released from domestic laundry wastewater [4]. Polyester is the largest component of MP fibres observed in effluent from commercial washing machines [5,6]. Household washing machines produce more than 1900 fibres during the washing of a single garment [7,8]. Polyester comprised a major component of plastics discharged from a selection of secondary and tertiary treatment plants in two different studies in Germany and the UK, indicating its importance in investigations of MP pollution within such systems [3,9]. Foams primarily consist of expanded polystyrene (PS), commonly used in construction for insulation purposes or in the food industry [10]. Microbeads typically originate from personal care products, which are also a primary source of MPs [11]. Tyre (rubber) wear particles (TWP) are also one of the main sources of MP pollution generated by the friction between tyres and roads and are transported into wastewater during rainfall [12]. Within municipal wastewater, antibiotics are also present and can interact with other contaminants in water such as MPs [13,14,15]. The adsorption of contaminants into the plastic surface transports them through treatment systems into the environment, across much larger distances that would otherwise be impossible [16]. MP pollution not only presents an environmental risk in itself, but is coupled with a potential increase in the spread of harmful organic contaminants which carry with them the potential to increase environmental antibacterial resistance [17].
Ceftazidime (CAZ) is a 3rd generation beta lactam cephalosporin antibiotic and is listed by the World Health Organization as a critically important antimicrobial [18]. CAZ-resistant bacteria have been identified in hospital and community wastewater and are therefore likely to be in contact with large numbers of MPs, also shown to be in such wastewater [19,20]. Some initial evidence supports this assumption, as in the presence of MPs, larger numbers of CAZ-resistant bacteria were identified in the intestinal communities of sea cucumbers [21]. Despite the importance of this drug to human health and the early evidence of bacteria resistance within wastewater, little information regarding the interaction with MPs is available in the literature. Due to the presence of CAZ alongside MPs in municipal wastewater, understanding their interaction is a key example of understanding how MPs can affect the fate of emerging pollutants within WWTPs via adsorption.
Data from the field of nanomaterials help us to understand the adsorption behaviour of CAZ, mainly from a remediation perspective. Multi-walled carbon nanotubes were tested as an adsorbent with a reported 70% removal in 10 min from a 30 mg L−1 solution of CAZ [22]. Removal reached a peak of 80% in 60 min, and the authors observed pH dependence linked to altered electrostatic interaction when CAZ was in the protonated state. Copper-based metal organic frameworks, doped-activated carbon, and fly ash zeolites have also been tested as high-capacity adsorbents for CAZ removal from water [23,24,25]. Although there are a few studies exploring the adsorption of antibiotics by weathered MPs (Table 1), the adsorption of CAZ by MPs has not been reported in the literature to our knowledge.
Table 1 shows that pristine MPs present very different adsorption properties to weathered MPs due to physical degradation and roughening of the surface as well as chemical oxidation leading to changes in hydrophobicity [40]. As MPs travel through WWTPs, they may be partially degraded by oxidative conditions, therefore changing their adsorption potential [14,41,42,43,44]. Moreover, MP and antibiotic degradation can take place, particularly in WWTPs where oxidative tertiary treatment systems are deployed. There is limited evidence of how these complex systems, containing large number of distinct pollutants, influence MP adsorption and the transportation of organic contaminants. In previous studies, pristine reference MPs of PE, PS, PVC, and PP were tested for the adsorption of organic contaminants, but the adsorption capacity of weathered MPs have not been widely tested yet [45]. This repulsive interaction will be reduced by MP ageing as the formation of oxygen containing functional groups by oxidation decreasing plastic hydrophobicity [46].
This paper presents the first study of CAZ adsorption by MPs with a focus on the changing behaviour of weathered MPs within WWTPs. We considered pristine and weathered samples of different polymer types with different sizes and shapes to build an understanding of complex MP antibiotic interactions within these systems.

2. Materials and Methods

2.1. Materials

Ten MPs of various polymer types and physical forms, described below (Table 2 and Figure S1), were considered. PET fibres with a diameter of 40 µm and length of 3 cm were attained from Phoenix Fibres, UK. Fibres were then cut into shorter MPs with a median length of 733 µm. PE beads were extracted from commercial facewash using a method modified from Cheung et al. [47]. Briefly, facewash was rinsed in deionised (DI) water pre-heated to 90 °C and stirred to dissolve the non-plastic solids. The solution was then filtered (2.4 µm glass fibre filter), and the MP beads rinsed with hexane to remove the final traces of soap and wax solids. The MPs were then rinsed a final time with deionised water and dried to a constant mass (50 °C).
Four commercial foams, i.e., hard/soft PS and PU, were bought from the market. Hard PS was procured from Styrodur 2800C, Bauhaus Europe. Soft PS was procured from FRAGMAT EPS F, Laško, Slovenia. Hard PS was procured from Styrodur 2800C, Bauhaus, and had a thickness of 20 mm. Soft PS was procured from FRAGMAT EPS F, having a thickness of 20 mm. Hard PU was procured from BACHL PIR MV, PIR/PUR 026, Bachl Kft., Tószeg, Hungary, having a thickness of 20 mm. Soft PU was procured from Repsol, Madrid, Spain.
Four tyres from various brands were also studied in this work with varied manufacturing dates. Bridgestone 225/45 R 17 91W (December 2016), Kumho 10.00 R 20 16PR (February 2002), Michelin 245/45 R 18 100V (April 2018), and Goodyear 205/55 R 16 91T (November 2014) (samples described in Supplementary Information). Ceftazidime (98%) was obtained from Cambridge Bioscience. Sodium hydroxide pellets and acetic acid (99.7%) were purchased from Fischer Scientific. Acetonitrile and Ultrapure water of HPLC grade were used in the HPLC analysis.

2.2. Artificial Weathering

Various weathering processes were applied in order to prepare MPs with a degradation degree relevant to wastewater systems. The weathering system was carefully selected in each case to provide relevant degradation levels representative of MPs in real systems. To prepare PET fibres with a controlled degree of degradation, a standardised hydrolysis treatment was applied which has been shown to reproducibly produce environmentally relevant reference materials [48]. In brief, fibres were treated for 3 h with a 10% sodium hydroxide solution at a temperature of 90 °C prior to rinsing with DI water and drying to constant mass (50 °C). When developing this method, Sarno et al. considered physical changes and also matched the ratio of terephthalic acid and ethylene glycol generated during hydrolysis to the UV degradation of fibres under natural conditions, showing the relevance of the prepared material to real environmental systems. PE beads were weathered for 9 h using an ultraviolet-activated hydrogen peroxide advanced oxidation method reported previously [49]. PS and PU foams were then weathered for 600 h in a Xenon test chamber, Q-SUN Xe-3 chamber, according to the ISO 4892-2:2013 [50] standard to replicate environmental degradation. During weathering, samples were exposed to 60 W m−2 of irradiation at 38 °C in the chamber with 50% relative humidity at a fan speed of 2000 rpm. TWPs were used as received from end-of-life commercial tyres and not subject to further artificial weathering. PS and PU foams and rubber tyre wear particles (TWPs) were cryogenically ball-milled in a Domel Tehtnica MillMix 20, Zelezniki, Slovenia.

2.3. Batch Adsorption Experiments

All adsorption experiments were conducted at room temperature (25 °C) and the initial solution pH was 6. MPs were weighed into glass vials, and 10 mL of a stock solution of CAZ in deionised water was added. The vials were capped and shaken on an orbital shaker (Denley orbital mixer platform) at a speed of 200 rpm. The initial concentration of CAZ was adjusted (1, 5, 10, and 15 mg L−1) in selected experiments and the typical MP concentration was 2 g L−1. At the end point, the solution was filtered using 0.22 µm syringe filters. Control experiments using identical parameters to adsorption conditions in the absence of antibiotics were conducted to measure the adsorption of CAZ onto the interior glass surface of the vial in the absence of MPs and in the absence of CAZ to consider the leaching of degradation products from the MPs into the water, which may have interfered with the HPLC analysis.

2.4. Analytical and Microscopy Techniques

The mass loss of plastic was used to track the weathering as reported by the scientific literature [37,40,51,52]. The MPs were weighed following a previously described protocol [49]. Scanning electron microscopy (SEM) was used to image the MP surfaces. MPs were secured onto conductive carbon tape and sputter-coated with gold. Images were collected on a JSM-IT800 Field Emission Scanning Electron Microscope in secondary electron imaging mode using an accelerating voltage of 15 kV and a 30 micron aperture at a working distance of approximately 10 mm. MP sizing was conducted using gathered SEM images and PE bead sizing was additionally performed using images gathered by optical microscopy where particles (n = 50) were sized using ImageJ software (https://imagej.net/ij/, accessed on 11 April 2025) and a micrometre scale. Atomic force microscopy (AFM) was used to further probe the roughness of the PET fibre samples. A Bruker/JPK NanoWizard 4 XP AFM, equipped with an Inverted Optical Microscope (Zeiss Axio Observer 5) was used. Imaging was carried out in tapping mode using NTESPA cantilever probes (Bruker AFM Probes, Camarillo, CA, USA) with a nominal spring constant and tip radius of 40 N/m and 8 nm, respectively. The data analysis was performed using Gwyddion software (http://gwyddion.net/, accessed on 9 April 2025) [53]. Images were sharpened, and a second-order polynomial background subtraction was applied to remove the influence of fibre curvature. The root mean square (RMS) of surface roughness was calculated from triplicate 5 × 5 µm images resulting in 3 RMS values for each sample tested. Fourier transform infrared spectroscopy (FTIR) was used to gather an IR spectra of each polymer surface to gain information on the presence of functional groups and chemical oxidation as a result of weathering [54]. FTIR spectra of the PET fibres and PE beads were collected using a PerkinElmer Spectrum Two Attenuated Total Reflection ATR-FTIR spectrometer. Spectra for the PU and PS foams were collected using a Perkin-Elmer Spectrum One ATR-FTIR. The carbonyl index (CI) was calculated as described previously [49,54].
The concentration of CAZ was determined using a Shimadzu HPLC with an SIL-20A HT auto sampler, LC-20AD Liquid chromatograph, CTO-20A column oven, and SPD-20A UV/VIS detector on a C18 column (250 × 4.6 mm, 5 mm) from Phenomenex, USA. Mobile phase acetonitrile (65%) and 0.1% acetic acid (45%), with a flow rate 1 mL min−1 and an injection volume of 10 µL were used. The peak area was compared to a standard curve to determine the concentration [55].
Adsorption Q (mg g−1) was calculated according to Equation (1) where C0 is the initial concentration of CAZ (mg g−1), Ce is the equilibrium concentration (mg L−1) of CAZ, m is the mass of adsorbent (g), and V is the total volume of the reaction mixture (L).
Q = C 0 − C e m × V

2.5. Adsorption Isotherms

A linear model and the Langmuir model were applied to calculate the adsorption isotherms of CAZ on PET MP fibres.
The linear model is described as follows:
Q = K d C e
The Langmuir isotherm model is described as follows:
Q = Q m a x K L C e 1 + K L C e
where Kd (L g−1) is the linear partition coefficient, Qmax is the maximum monolayer adsorption capacity (mg g−1), and KL (L mg−1) is a coefficient describing the adsorption interaction between the plastic surface and CAZ.

3. Results

3.1. Adsorption of Ceftazidime by Microplastics

In wastewater treatment systems, a broad range of MP shapes, types, and degradation degrees are present, impacting their interaction with co-contaminants [39,56]. The adsorption of antibiotics by each MP will vary, thereby complicating attempts to quantify their spread in real systems. Figure 1 shows the SEM images of representative MPs potentially found in WWTPs that were tested for CAZ adsorption. Additional photos (Figure S1) and images to show hard and soft variants of PS and PU and all tyre brands are included in the Supplementary Information (Figures S4 and S5).
The MPs selected for this study showed diverse particle sizes and surface morphologies. PET (Figure 1a) is semicrystalline and polar with aromatic groups and oxygen containing functionalities providing the potential for intermolecular interactions with CAZ molecules in water. The PET fibres displayed a very smooth surface in SEM imaging with few features evident. In contrast, PE beads (Figure 1b) extracted from facewash had a very rough surface, likely due to their intended function as exfoliants. The extracted PE beads varied in size between 250 and 850 µm as shown in Figure S2. Cracked and fragmented beads were observed in all assayed samples as well as whole spheres, matching the observations of MPs in facewash samples in the literature [57]. PS foam (Figure 1c) is characterised by a flaked structure with a large number of small fragments collected in agglomerations. PU foam MPs (Figure 1d) are composed of small (50–80 μm diameter) aggregations of many small MPs with the individual fragments having a granular appearance. TWPs (Figure 1e) are sized within the range of 300–1000 µm and exhibit a rough, textured surface. The adsorption of CAZ by these MPs was studied over a 24 h period and the results are shown in Figure 2.
As a preamble, it was found that CAZ adsorption by the assayed types of MPs ranged between 0.15 mg g−1 and 1.432 mg g−1, which is comparable with results of previous studies (shown in Table 1). For example, the chlortetracycline antibiotic was adsorbed at 0.72 mg g−1 by pristine PE, ciprofloxacin was adsorbed at 0.15 mg g−1 by pristine PS, and tetracycline was adsorbed at 0.917 mg g−1 by weathered PE. In Figure 2a, it is observed that pristine PET fibres reached a maximum adsorption of 0.664 mg g−1 within the first 3 h of contact, which was then maintained almost unchanged throughout the rest of the contact time. However, the CAZ adsorption rate by the weathered PET fibres was initially slower than that of the pristine ones, but after 24 h, the weathered fibres reached the highest adsorption capacity of 0.773 mg g−1 compared to all the other MPs. Interestingly, the CAZ adsorption capacity kept increasing, reaching 1.432 mg g−1 after 48 h of contact time. Figure 2b shows that the adsorption of CAZ by PE MP beads increased linearly with time for both the raw extracted and weathered beads, reaching 0.210 and 0.234 mg g−1 in 24 h, respectively. The adsorption capacity was higher for the weathered beads compared with that of the raw extracted ones. Regarding the soft PS foams displayed in Figure 2c, there was a substantially larger capacity for CAZ adsorption than the hard PS samples, reaching 0.53 mg g −1 and 0.48 mg g −1 in 24 h for pristine and weathered soft PS samples, respectively, but the CAZ adsorption for hard PS was only 0.24 mg g−1 (weathered) and 0.17 mg g−1 (pristine) after 24 h. This can be explained by the fact that high-density polymers have slower mass transfer at the plastic surface, a finding that is in agreement with a previous study measuring tetracycline adsorption, during which low-density polyethylene was found to have a greater adsorption capacity than high-density polyethylene due to mass transfer differences [46]. The difference in the adsorption capacity of soft and hard PS foams could also be influenced by the presence of additives, which may inhibit the accumulation of other organics on the PS surface. Weathered foams were tested for adsorption of CAZ, and both hard and soft PS showed almost unchanged uptakes over 24 h when compared to the pristine samples. It should be noted though that adsorption for weathered soft PS seems to take place slightly faster (at about 5 h) compared to that for pristine soft PS, when an equilibrium was reached after about 12 h. Soft and hard PU foams had a low overall uptake, reaching 0.17 mg g−1 and 0.19 mg g−1, respectively, after 24 h (Figure 2d). It was also observed that the initial uptake increased more slowly for the hard PU compared to that for the soft PU foam, again related to the rate of mass transfer affected by the extent of chain branching and crosslinked polymer structures that can inhibit the movement of organics within the structure [58]. Weathered PU adsorption could not be determined by HPLC due to interference from degradation products which could not be readily separated from CAZ. Figure 2e demonstrates that the Kumho TWPs had the highest adsorption capacity of 0.495 mg g−1 after 24 h, followed by those of Goodyear (0.288 mg g−1), Michelin (0.273 mg g−1), and Bridgestone (0.212 mg g−1). This difference in maximum value can be rationalised when the differences of each rubber are considered. Kumho TWPs are natural rubber whereas Bridgestone, Michelin, and Goodyear TWPs are artificial rubbers manufactured for passenger vehicle tyres made with styrene-butadiene rubber (SBR), which may influence the adsorption behaviour due to their differing properties [59]. Also, the Kumho tyres were manufactured 12 years before the earliest passenger tyre sample (Goodyear 2014), indicating that during its useful life, there may be a higher degree of natural weathering, leading to changes in the adsorption properties [14]. Interestingly, Goodyear and Michelin showed rapid increases in CAZ adsorption in the first 3 h, while Bridgestone increased at a slower linear rate, reaching maximum adsorption after 24 h.
The results shown in Figure 2 indicate that the dominant factor influencing CAZ adsorption on MPs was weathering due to its impact on surface properties and chemical reactivity. In the present study, the adsorption of CAZ on different polymer types, sizes, and shapes, as well as the effect of weathering, is discussed. Pristine and weathered PET, PE, and PS were studied for CAZ adsorption. Weathered PET and PE in particular showed high CAZ adsorption, whereas weathered PS exhibited adsorption similar to that of pristine PS. The textured surface of weathered MPs, due to cracks and fragments, promotes the adsorption of organic contaminants, as demonstrated in the literature, where a 5-fold increase in tetrabromobisphenol adsorption was recorded for cracked PE beads as compared to reference PE spheres of the same size [60]. Furthermore, weathering increases adsorption capacity by increasing surface area through physical degradation. Fan et al. demonstrated this effect by showing that TWPs had a greater adsorption capacity than PE MPs for chlortetracycline and amoxicillin [12]. The almost unchanged CAZ uptake by pristine and weathered PS aligns with the results presented by Wu et al., where aged PS was shown to have a reduced capacity for adsorbing a polybrominated diphenyl ether due to a decrease in the number of available adsorption sites [55].
Moreover, Figure 2 shows that the type of MPs affects CAZ adsorption, and PET fibres yielded the highest adsorption capacity. The most common mechanism by which MPs adsorb organic pollutants is hydrophobic interactions. Additionally, other mechanisms, such as electrostatic interactions and non-covalent forces like hydrogen bonds, halogen bonds, and π-π interactions, also contribute to the adsorption of pollutants into MPs. For example, π-π interactions are possible when the polymer chain contains aromatic groups, explaining the higher adsorption of PET and PS compared to PE. This trend is supported by Li et al., who demonstrated that PS had a greater adsorption capacity than PE for antibiotics due to these interactions [56]. Functional groups within the polymer chain, such as those in PU, may also form hydrogen bonds with antibiotics [61], although the low PU adsorption of CAZ observed in our study indicates that the contribution of this interaction is not significant.
In addition to polymer type, additives and functional groups also impact adsorption. The study examined three passenger tyre wear particle (TWP) samples, made from styrene-butadiene rubber, and one truck TWP sample, made from natural rubber. The truck TWP sample exhibited twice the adsorption capacity of the passenger TWPs, confirming the role of functional groups in CAZ adsorption.
Size and shape contributions to adsorption capacity are less critical factors. Smaller MPs have a higher surface-area-to-volume ratio, enhancing adsorption, while irregularly shaped MPs, such as fibres and fragments, provide more surface roughness and binding sites than smooth spheres. This explains the lower adsorption of PE beads compared to PET fibres, PS foams, and TWP fragments.
Overall, CAZ adsorption followed the following order (from highest to lowest capacity): weathered PET fibres > pristine PET fibres > soft PS (both pristine and weathered) > Kumho TWPs > weathered soft PS > Goodyear TWPs > Michelin TWPs > weathered hard PS > weathered PE > Bridgestone TWPs > pristine PE > pristine hard PU > pristine soft PU = pristine hard PS. In general, it was observed that weathering increases the adsorption of ceftazidime which can be explained by considering the increase in surface area by the physical degradation of MPs during weathering. The chemistry and structure of the MPs assayed, with an emphasis on PET fibres, which had the highest adsorption capacity, are analysed in Section 3.2 and Section 3.3 to better understand the role of weathering and surface modifications in adsorption capacity.

3.2. The Effect of Weathering on Microplastic Surface Structure

SEM and AFM analyses of the weathered PET MPs, which yielded the highest adsorption capacity, were carried out to study the effect of ageing on the structure of the polymer surface. The results are presented in Figure 3, Figure 4 and Figures S3–S5. The weathered PET fibres showed clear changes on their surface imaged by SEM after the weathering process, as shown in Figure 3.
Cracking and fragmentation features were visible for the weathered fibres, as was surface hole and dimple formation. This weathering is consistent with previously observed accelerated fibre degradation under advanced oxidation treatment in wastewater and hydrolysis treatment [48,49]. It should be also noted that the weathered fibres were greatly reduced in bulk mass by 74% compared to the pristine fibres, indicating the aggressiveness of the hydrolysis weathering process that was applied and the high deterioration of the surface of the fibres.
Therefore, the higher maximum capacity of weathered PET can be explained by a greatly increased surface area, as shown by the increased surface cracking and pitting. The change in the initial rate of CAZ adsorption may have been influenced by the presence of MP degradation products in the solution competing for adsorption sites or interactions with the antibiotic itself. Similarly, the competition for adsorption sites was shown to reduce the rate of antibiotic adsorption as the MP size was decreased to the nanoscale [39]. During the weathering process of PET fibres, particles are fragmented into small micro- and nano-fragments, which may also affect the rate of adsorption [49].
As weathered PET fibres displayed the highest adsorption capacity of all the MPs tested, the surface morphology was further investigated at the nanoscale. The surface roughness was measured by AFM, and the representative height plots and 3D surface maps for the pristine and weathered fibre surfaces are shown in Figure 4a,b. It can be seen that holes and ridges are visible in the images as dark and white areas of the 2D plot, respectively. As compared to the untreated PET fibre, a large increase in surface roughness was observed with the mean value increasing from 9.23 nm (pristine fibres) to 31.38 nm after weathering (Figure 4c). This increase in the roughness of the surface at both the nano- (Figure 4) and micro- (Figure 3) scale is expected to positively influence the adsorption capacity of the MPs.
The surfaces of all the MPs were imaged by SEM (supplementary information, Figures S3–S5). Upon the weathering of PE beads, a trend towards a smaller diameter from a starting mean value of 552 µm to 485 µm was observed (Figure S2), although the overall range of bead size remained the same. Understanding their fragmentation during weathering sheds light on the potential change in toxicity. PE MPs have been shown to be size-dependent due to an increased retention of small fragments in biota [62,63]. This was accompanied by a bulk mass loss of 10.8% over 9 h of weathering. The trend in bead size towards smaller-diameter MPs after weathering will affect the adsorption behaviour. Smaller particle sizes are expected to accompany increased adsorption capacity due to surface area increases. However, competition at the surface was reported at a threshold size in the adsorption of ciprofloxacin onto PS micro- and nano-plastics [39], indicating that very small particle sizes may impede the rate of antibiotic adsorption. Significant changes to PS and PU foams were not observed in the SEM images, which explains the limited change in maximum adsorption between pristine and weathered samples of each of these polymer types. Without a large change to the physical surface of these MPs, weathering did not result in significant changes to the adsorption.

3.3. The Effect of Weathering on Polymer Surface Chemistry

The weathering of polymer surfaces typically introduces oxygen containing functional groups, which may be capable of stronger interactions with antibiotics, thus altering the adsorption behaviour of MPs [27,64]. FTIR spectra were gathered for PET, PE, PS, and PU samples before and after weathering with the results summarised in Table 3. Where possible, the CI was calculated from the FTIR spectra.
PET fibres showed a decrease in CI during hydrolysis weathering. This is in contrast to samples aged by UV-based weathering methods, where the CI increases with time during weathering [49]. The release of carbonyl groups in carboxylic acid degradation products by the hydrolysis of ester bonds in the polymer is responsible [48,65]. PE beads have an averaged FTIR spectra unchanged by the weathering process, with no emergence of a clear carbonyl signal. Compared to a sample of PE film, the spectra of PE beads showed typical identifying peaks for linear low-density polyethylene [66]. The variable weathering of individual beads may not be captured in these data due to the FTIR spectra being measured as an average signal from a number of PE beads or an insufficient weathering time to observe significant chemical changes. To investigate this, a micro-FTIR analysis of individual PE beads displaying the physical characteristics of heavy weathering as observed in the SEM images may be a useful avenue for future work to show the localised formation of oxygen containing groups [67]. For PS and PU foams, the weathering of the MPs increased the CI.

3.4. PET Fibre Microplastic Adsorption Isotherms

The adsorption isotherms were calculated for pristine and weathered PET fibres, which presented the highest CAZ adsorption capacity. Figure 5 shows the adsorption capacity in relation to the CAZ concentration in water while the isotherm parameters are listed in Table 4. The plotted isotherm curves are included in the supplementary information (Figure S6, Table S1).
The concentration of CAZ in solution and the concentration of adsorbent (MPs) were varied, and the adsorption isotherms were plotted according to the linear and Langmuir models described in Section 2.5. These models were selected to enable comparison across other systems as they are commonly used in literature studies considering organic contaminant adsorption into MPs [12,61,64,68].
The adsorption capacity of PET fibres for CAZ varied significantly between the pristine and weathered samples. Both isotherm models showed good fitting for data in the CAZ concentration range of 1–12.5 mg L−1. The adsorption of ceftazidime into the MP surface is well described by the Langmuir isotherm, indicating a monolayer process. This agrees with observations made by Sun et al. [61] who observed the monolayer adsorption of norfloxacin into PS and PE MPs. Surface area, crystallinity, and surface charge are all factors likely to contribute to the change between pristine and weathered samples [69]. No reported adsorption data for CAZ on MPs currently exist in literature. The reported Kd values from a linear model for sulfamethoxazole and sulfamethazine uptake by virgin PET were 0.0222 and 0.0226 L g−1 [68,70], which were similar to the Kd obtained for CAZ adsorption in this study, while there were no literature values available for the adsorption of antibiotics on weathered PET [64].

4. Conclusions

The effect of MP type, shape, and ageing on the adsorption of CAZ antibiotic by MPs was studied via lab scale experiments. Varied rates of uptake and maximum adsorption capacity were linked to the polymer types, highlighting the importance to consider the specific MPs present in wastewater when studying interactions with chemical co-contaminants. Five polymers (PET, PE, PS, PU, and TWPs) in different forms (fibres, beads, foam and fragments) and sizes (10–1000 µm) were compared. In most cases, the weathering of MPs increased the surface roughness and altered the chemical properties as measured by SEM, AFM, and FTIR. The weathered PET fibres exhibited the highest capacity for CAZ adsorption, reaching 1.432 mg g−1 after 48 h of contact, as compared to 0.664 mg g−1 for pristine PET fibres on the same timescale. The increased adsorption capacity, compared to those of the other MPs assayed, was linked to the larger surface area, created by the formation of cracks, holes, and pitting formed during the weathering process. This trend was not extended across all MP samples, with PE and PS showing negligible increases in adsorption upon weathering and the overall order as follows: PET fibres (weathered) > PET fibres (pristine) > soft PS (pristine) > Kumho TWPs > soft PS (weathered) > Goodyear TWPs > Michelin TWPs > hard PS (weathered) > PE weathered > Bridgestone TWPs > PE (pristine) > hard PU (pristine) > soft PU (pristine) = hard PS (pristine). The adsorption of antibiotics by MPs within wastewater increased the potential for the transportation of pollution within the environment. The MP type, shape, and ageing will heavily influence the rate and capacity of antibiotic adsorption and must be carefully considered in order to understand their impact on the aquatic environment.

Supplementary Materials

The following supporting information can be downloaded at: https://www.mdpi.com/article/10.3390/environments12040120/s1, Figure S1: Photos of MPs types studied for ceftazidime adsorption. Polyurethane (PU); Polystyrene (PS); Polyethylene (PE); polyester (PET) and tyre wear particles (TWP); Figure S2: Size distribution of PE beads extracted (pristine) from facewash and weathered by advanced oxidation process; Figure S3: SEM images of PE beads extracted from facewash before (top) and after (bottom) weathering; Figure S4: SEM images of (a) hard PS (b) soft PS (c) hard PU and (d) soft PU before (left) and after (right) weathering; Figure S5: SEM images of (a) Bridgestone, (b) Kumho, (c) Michelin, (d) Goodyear TWPs; Figure S6: Isotherm fitting plots for linear (left) and Langmuir (right) models; Table S1: Isotherm fitting data for ceftazidime adsorption by PET fibres.

Author Contributions

Conceptualization, T.E. and E.C.; methodology, T.E., V.B. and A.A.; software, T.E.; validation, Y.H., Q.Z. and V.B.; formal analysis, T.E. and V.B.; investigation, T.E and V.B.; resources, E.C., V.K. and A.A.; data curation, Y.H. and Q.Z.; writing—original draft preparation, T.E. and V.B.; writing—review and editing, E.C., V.K. and A.A.; supervision, E.C. and V.K.; funding acquisition, T.E. and V.B. All authors have read and agreed to the published version of the manuscript.

Funding

This work was supported by a NERC Doctoral Training Partnership grant (NE/S007407/1). This project received funding from the European Union’s Horizon 2020 research and innovation program under Marie Sklodowska-Curie, grant agreement No. 860720 and COST Action PRIORITY CA20101.

Data Availability Statement

The original contributions presented in this study are included in the article/Supplementary Material. Further inquiries can be directed to the corresponding author.

Acknowledgments

For the purpose of open access, the authors have applied a Creative Commons Attribution (CC BY) licence to any Author-Accepted Manuscript version arising from this submission.

Conflicts of Interest

The authors declare no conflicts of interest.

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Environments 12 00120 g001
Figure 1. Representative SEM images of (a) pristine PET, (b) as-extracted PE, (c) pristine PS and (d) pristine PU, and (e) as-extracted TWP MPs (left) and their magnified surface structure (right).
Figure 1. Representative SEM images of (a) pristine PET, (b) as-extracted PE, (c) pristine PS and (d) pristine PU, and (e) as-extracted TWP MPs (left) and their magnified surface structure (right).
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Environments 12 00120 g002
Figure 2. Adsorption capacity of CAZ over time by (a) PET fibres, (b) PE beads, (c) PS foams, (d) PU foams, and (e) TWP rubbers (MPs concentration 2 g L−1, CAZ concentration 10 mg L−1).
Figure 2. Adsorption capacity of CAZ over time by (a) PET fibres, (b) PE beads, (c) PS foams, (d) PU foams, and (e) TWP rubbers (MPs concentration 2 g L−1, CAZ concentration 10 mg L−1).
Environments 12 00120 g002
Environments 12 00120 g003
Figure 3. Representative SEM images of weathered PET fibres (NaOH alkaline hydrolysis, 3 h) (a,b) are different images of fibres samples.
Figure 3. Representative SEM images of weathered PET fibres (NaOH alkaline hydrolysis, 3 h) (a,b) are different images of fibres samples.
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Environments 12 00120 g004
Figure 4. AFM height images and surface maps of (a) pristine PET fibre, (b) weathered PET fibre, and (c) roughness (RRMS) of pristine and weathered polyester fibres (mean values marked, 25/75th percentile and standard deviation represented in the box and whisker plot).
Figure 4. AFM height images and surface maps of (a) pristine PET fibre, (b) weathered PET fibre, and (c) roughness (RRMS) of pristine and weathered polyester fibres (mean values marked, 25/75th percentile and standard deviation represented in the box and whisker plot).
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Environments 12 00120 g005
Figure 5. CAZ adsorption by PET fibre MPs (working volume 10 mL, MP concentration 2 g L−1).
Figure 5. CAZ adsorption by PET fibre MPs (working volume 10 mL, MP concentration 2 g L−1).
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Table 1. Studies on antibiotic adsorption on weathered microplastics.
Table 1. Studies on antibiotic adsorption on weathered microplastics.
AntibioticsMPs TypeMPs SizeAdsorption CapacityWeathering MethodRef.
SulphanilamidePolyamide (PA),
Polyvinyl chloride (PVC) and PET		150 µm19.70 μg/g (W)
24.27 μg/g (W)
12.45 μg/g (W)		Ultraviolet (UV) irradiation and temperature fluctuation[26]
Ciprofloxacin (CIP)Polystyrene (PS)-2.16 mg/g (P)
5.45 mg/g (W)		Fenton immersion method[27]
Chlortetracycline (CTC) and Amoxicillin (AMX)TWP and
PE		≤74 μm (TWP)CTC: 0.90 mg/g (P); 2.09 mg/g (W)
AMX: 2.25 mg/g (P); 2.91 mg/g (W)
CTC: 0.72 mg/g (P); 1.57 mg/g (W)
AMX: 1.65 mg/g (P); 2.72 mg/g (W)		Heat-activated potassium persulfate[12]
Tetracycline (TC) and CIPPolylactic acid (PLA) and
PVC		250–550 μm
75–150 μm		TC: 0.96 mg/g (P); 1.57 mg/g (W)
CIP: 0.67 mg/g (P); 0.85 mg/g (W)
TC: 2.51 mg/g (P) to 5.49 mg/g (W)
CIP: 3.19 mg/g (P); 3.77 mg/g (W)		UV irradiation[28]
CIPPS
PVC		∼75 μmKinetic model at 10 mg/L (PS < PVC < Aged PVC < Aged PS)UV-accelerated aging[29]
Oxytetracycline (OTC)PLA75–150 μm581 μg/g (P); 1193 μg/g (W);
728 μg/g (Biofilm)		Incubated in sewage influents[30]
Cu(II) and OTCThermoplastic Polyurethane (TPU)50–100 μmCu(II): 0.613 mg/g (P); 0.576 mg/g (W)
OTC: 0.980 mg/g (P); 1.506 mg/g (W)		UV irradiation[31]
Benzoic acid (BA), Sulfamethoxazole (MX), Sulphamerazine (MR) and CIPPA75–150 μmBA (0.75 mg/g) > SMX (0.54 mg/g) > SMR (0.11 mg/g) > CIP (0.08 mg/g))Xenon lamp irradiation, H2O2, and heat-activated K2S2O8 weathering[32]
TrihalomethanesPVC125–300 µm10–20 µg/g (P)
Reduced by 10% (W)		Metal halide lamp[33]
TC, Levofloxacin (OFL), CIPPE, PS, Poly(butylene adipate-co-terephthalate) (PBAT)30–500 μm (PE), 75 μm (PS), and 150 μm (PBAT)PS for TC, CIP increased by 58% (0.69–1.09 mg g−1), 138% (1.72–4.09 mg g−1);
PBAT for CIP increased from 0.62 mg g−1 to 6.22 mg g−1
PE for OFL is 0.35 mg g−1		UV weathering[34]
OTCPS0.45–1 mmFreundlich Kf value of 894 ± 84 ((mg kg−1) (mg L−1)1/n) for beached foamPlastic debris from coastal beaches[35]
TCPE150–250 µm120.5 μg/g (P)
91.7 μg/g (W) at pH = 7, 20 °C; 111.1 μg/g (W) at pH = 4; 114.9 μg/g (W) at pH = 10; 120.5 μg/g (W) at 2 mg/L HA; 48.1 μg/g (W) at 10 mg/L HA		Aged at different pH, temperature, ionic strength, ageing time, and humic acid[36]
CIPPS and
High-density polyethylene		50.4 ± 11.9 μm
45.5 ± 12.9 μm		71.86 mg/g (P); 251.59 mg/g (W) for K2S2O8; 189.52 mg/g (W) for Fenton treatment.
51.61 mg/g (P); 167.19 mg/g (W) for K2S2O8; 126.21 mg/g (W) for Fenton treatment		Heat-activated K2S2O8 and Fenton treatments[37]
Penicillin (PNC)PS80 nmPercentages of PNC adsorption on 73.0 ± 1.3 (PS); 80.8 ± 4.3 (UV), and 69.1 ± 6.4 (Ozonation)UV irradiation and ozonation[38]
CIP and bisphenol-A (BPA)PS∼40 nm0.15 mg/g (P); 4.07 mg/g (W) for CIP
4.92 mg/g (P); 8.71 mg/g (W) for BPA		UV irradiation ageing[39]
P: Pristine; W: Weathered.
Table 2. Microplastic samples studied for CAZ adsorption.
Table 2. Microplastic samples studied for CAZ adsorption.
Microplastic TypeSize Range (µm)ShapeArtificial Weathering Process
Polyester (PET)40 (diameter)
70–2117 (length)		FibreAlkaline hydrolysis (10% NaOH. 3 h)
Polyethylene (PE)250–850BeadsUVC-activated H2O2 (500 mg L−1, 9 h)
Hard polystyrene (Hard PS)50–80Foam particleXenon lamp (600 h)
Soft polystyrene (Soft PS)50–80Foam particleXenon lamp (600 h)
Hard polyurethane (Hard PU)10–80Foam particleXenon lamp (600 h)
Soft polyurethane (Soft PU)10–80Foam particleXenon lamp (600 h)
Bridgestone Tyre300–1000Rubber fragment-
Kumho Tyre300–1000Rubber fragment-
Michelin Tyre300–1000Rubber fragment-
Goodyear Tyre300–1000Rubber fragment-
Table 3. Calculated carbonyl index of PET, PS, and PU MP samples.
Table 3. Calculated carbonyl index of PET, PS, and PU MP samples.
Carbonyl Index (CI)
PristineWeathered
PET6.75.5
Hard PS0.312.35
Soft PS0.373.51
Hard PU2.262.39
Soft PU0.741.70
Table 4. Linear and Langmuir isotherm fitting parameters for CAZ (1–12.5 mg L−1) adsorption by pristine and weathered PET fibres.
Table 4. Linear and Langmuir isotherm fitting parameters for CAZ (1–12.5 mg L−1) adsorption by pristine and weathered PET fibres.
PET FibersLinearLangmuir
Kd (L g−1)r2Qmax (mg g−1)KL (L mg−1)r2
Pristine0.023 ± 0.0020.9831.10 ± 0.090.5960.987
Weathered 0.045 ± 0.0060.9801.27 ± 0.090.7530.995
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Easton, T.; Budhiraja, V.; He, Y.; Zhang, Q.; Arora, A.; Koutsos, V.; Chatzisymeon, E. Antibiotic Adsorption by Microplastics: Effect of Weathering, Polymer Type, Size, and Shape. Environments 2025, 12, 120. https://doi.org/10.3390/environments12040120

AMA Style

Easton T, Budhiraja V, He Y, Zhang Q, Arora A, Koutsos V, Chatzisymeon E. Antibiotic Adsorption by Microplastics: Effect of Weathering, Polymer Type, Size, and Shape. Environments. 2025; 12(4):120. https://doi.org/10.3390/environments12040120

Chicago/Turabian Style

Easton, Thomas, Vaibhav Budhiraja, Yuanzhe He, Qi Zhang, Ayushi Arora, Vasileios Koutsos, and Efthalia Chatzisymeon. 2025. "Antibiotic Adsorption by Microplastics: Effect of Weathering, Polymer Type, Size, and Shape" Environments 12, no. 4: 120. https://doi.org/10.3390/environments12040120

APA Style

Easton, T., Budhiraja, V., He, Y., Zhang, Q., Arora, A., Koutsos, V., & Chatzisymeon, E. (2025). Antibiotic Adsorption by Microplastics: Effect of Weathering, Polymer Type, Size, and Shape. Environments, 12(4), 120. https://doi.org/10.3390/environments12040120

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