# UV-degradation is a key driver of the fate and impacts of marine plastics. How can laboratory experiments be designed to effectively inform risk assessment?

Laura M Hernandez, Lucy Howarth-Forster, Lisbet Sørensen, Andy M Booth, Alice Vidal, Nathalie Tufenkji, Richard Sempéré, Natascha Schmidt

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Laura M Hernandez, Lucy Howarth-Forster, Lisbet Sørensen, Andy M Booth, Alice Vidal, et al. UV-degradation is a key driver of the fate and impacts of marine plastics. How can laboratory experiments be designed to effectively inform risk assessment?. Marine Pollution Bulletin, 2025, 219, pp.118271. 10.1016/j.marpolbul.2025.118271. hal-05108475

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Submitted on 11 Jun 2025

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# UV-degradation is a key driver of the fate and impacts of marine plastics. How can laboratory experiments be designed to effectively inform risk assessment?

Laura M. Hernandez<sup>a</sup>, Lucy Howarth-Forster<sup>b</sup>, Lisbet Sørensen<sup>c,d</sup>, Andy M. Booth<sup>c</sup>, Alice Vidal<sup>e</sup>, Nathalie Tufenkji<sup>a</sup>, Richard Sempéré<sup>f</sup>, Natascha Schmidt<sup>g,*</sup>

## ARTICLE INFO

### Keywords:

- Plastic fragmentation
- Microplastic
- Marine litter
- Sunlight
- Seawater
- Plastic pollution

## ABSTRACT

Marine plastic litter is subject to different abiotic and biotic forces that lead to its degradation, the main driver being UV-induced photodegradation. Since UV-exposure leads to both physical and chemical degradation of plastic, leading to a release of micro- and nanoplasics as well as leaching of chemicals and degradation products – it is expected to have radical impacts on plastics fate and effects in the marine environment. The number of laboratory studies investigating the mechanisms of plastic UV-degradation in seawater has increased significantly in the past 10 years, but are the exposures designed in a manner that allow observations to be extrapolated to environmental fate? Most studies to date focus on quantifying plastic fragmentation and surface changes, but is this relevant for impact assessments? Here, we provide a review of the current scientific literature on UV-degradation of plastic under marine conditions. Plastic fragmentation processes and surface changes as well as implications of UV-degradation of plastics on additive leaching and the toxicity of UV-weathered versus non-weathered plastics are highlighted. Furthermore, experimental set-ups are critically inspected and recommendations for future studies are issued.

## 1. Introduction

While its durability makes plastic attractive to both manufacturers and consumers by promising a long product lifetime, the same material properties present a persistence problem once plastic leaks from the value chain and enters the environment. As for other pollutants, the marine environment is considered the final sink for a large proportion of leaked plastic and so the persistence and the mechanisms of plastic degradation are particularly important to understand from an environmental risk and impact perspective (GESAMP, 2015). Most plastics, including thermoplastics, will degrade to some extent in the natural environment through a range of abiotic and biotic processes, including thermo-oxidative degradation, biodegradation, photodegradation and hydrolysis. However, the degradation rates are influenced by a number

of intrinsic parameters (e.g., polymer type, surface area, additive chemical content) and extrinsic factors (e.g., exposure to varying levels of ultraviolet radiation, temperature and energy, as well as microbial community composition). The result of macroplastic litter degradation is the release of microplastic particles and plastic chemicals into the surrounding environment (Paluselli et al., 2019; Gewert et al., 2015; Sørensen et al., 2021; Fauvelle et al., 2021).

Degradation via ultraviolet (UV) irradiation-induced oxidation (photodegradation or photooxidation) – hereafter called UV-degradation – is the environmental mechanism regarded as most effective at degrading conventional (petroleum-based) thermoplastics such as polyethylene (PE), polypropylene (PP), polystyrene (PS), polyethylene terephthalate (PET) (Gewert et al., 2015). UV-B wavelengths (290–315 nm) are hereby significantly more effective in degrading

<footnote><sup>a</sup> Department of Chemical Engineering, McGill University, Montreal, Quebec, Canada, H3A 0C5</footnote>

<footnote><sup>b</sup> School of Geography, Earth and Environmental Sciences, University of Plymouth, Drakes Circus, Plymouth, PL4 8AA, UK</footnote>

<footnote><sup>c</sup> SINTEF Ocean AS, Department of Climate and Environment, Trondheim, Norway</footnote>

<footnote><sup>d</sup> Norwegian University of Science and Technology (NTNU), Department of Chemistry, Trondheim, Norway</footnote>

<footnote><sup>e</sup> Aix Marseille Univ., Toulon Univ., CNRS, IRD, Mediterranean Institute of Oceanography (MIO) UM 110, Marseille, France</footnote>

<footnote><sup>f</sup> Aix-Marseille Université, CNRS, LCE, UMR 7376, Ocean Science Institute, Marseille, France</footnote>

<footnote><sup>g</sup> NILU, Tromsø, Norway</footnote>

<footnote>* Corresponding author at: NILU, Fram Centre, Hjalmar Johansens gate 14, 9007 Tromsø, Norway. E-mail address: nsch@nilu.no (N. Schmidt).</footnote>
conventional thermoplastics than UV-A wavelengths (315–400 nm) or visible light (400–700 nm). In the marine environment, biofouling and the limited penetration of UV-light into the water column limit UV-degradation of plastics compared to e.g. beach environments (Jansen et al., 2024). UV-degradation also facilitates other degradation mechanisms via causing chain scission of the polymers as well as eventual fragmentation of the bulk material (Gewert et al., 2015). During the exposure of larger plastic items to UV radiation, fragmentation processes can lead to the formation of micro- and nanoplastics (Lambert and Wagner, 2016), while microplastics will further fragment to smaller particles (Sait et al., 2021).

Changes resulting from UV-degradation concern not only surface characteristics (e.g., surface cracking), but also the chemical composition of the plastic (Ainali et al., 2021; Fan et al., 2021; Mao et al., 2020). Most plastics contain a variety of chemicals, intentionally added or non-intentionally added substances (NIAS), that are typically not bonded directly to the polymer matrix (Wagner et al., 2024). Plastic additives of organic (e.g., phthalates, bisphenols, organophosphate esters) or inorganic nature (e.g., zinc or aluminum composites) are incorporated into the polymer matrix to impart certain properties (e.g., flexibility, color, flame-retardance, oxidation prevention, UV protection, heat- and microbial resistance) during the manufacturing of plastics or in the final product (Hahladakis et al., 2018). NIAS include the residues of production chemicals such as monomers, catalysts, reagents, and by-products, as well as degradation products and chemicals sorbed during the life cycle of the plastic. Additive chemicals and NIAS can be released into the surrounding environment during use of the plastic product or during post-use disposal and subsequent degradation (Hahladakis et al., 2018). By causing fragmentation and increasing the available surface area, it has been shown that UV-degradation plays a significant role in determining the degree of leaching and composition of plastic leachates and thus ultimately the impact of plastics on marine life through release of toxic chemicals and degradation products (Fig. 1) (Sørensen et al., 2021; Sait et al., 2021).

It is evident that UV-degradation can strongly influence the environmental impact of plastic that ends up in marine waters. For example, the appearance of holes on the material surface can lead to increased biofouling or trapping of particles such as sand, subsequently leading to changes in density and sinking rates (Zheng et al., 2023). A robust understanding of UV-degradation mechanisms and degradation rates, as well as formation and lifetime of microparticles and chemicals released via UV-degradation, is important for improved hazard and risk assessment. Furthermore, understanding how plastic fate and toxicity is influenced by UV-degradation is of particular interest in the context of global climate change. With the gradual recovery of the stratospheric ozone that is currently underway, future changes on solar UV radiation will depend on factors which are directly affected by climate change, for instance the formation of aerosols of solid and liquid particles which reduces and scatters UV radiation. Another important variable is the type and amounts of clouds formed, which is also changing locally. Finally, the melting of icecaps and the reduction in overall snowfall alter the Earth's surface reflectivity, exposing more soil and water that were previously covered by snow or ice. (Barnes et al., 2019). To this end, we are dependent on well-designed laboratory protocols and study designs that provide reliable data that can be extrapolated to real scenarios. There are currently no standardized protocols designed specifically to assess the production of particulate and chemical degradation products resulting from UV exposure of thermoplastics, nor any general consensus on how such studies should be set up. This means that it is likely that UV-degradation data produced in different studies is difficult to compare directly and challenging to extrapolate from. Furthermore, it remains challenging to accurately recreate environmentally relevant UV degradation exposure conditions in the laboratory, further impacted by the slow degradation rate of plastics and in relation to the timeframes commonly foreseen to conduct experiments.

Here, we provide a critical review of the most recent literature on the

UV-degradation of thermoplastics in marine environments. Particular focus is given to physicochemical surface changes and fragmentation mechanisms leading to the formation of microplastics. We also assess current knowledge regarding the release of chemical additives, NIAS and molecular degradation products during UV-degradation, as well as presenting a state-of-the-science on the toxicological implications of partially degraded plastic and associated chemicals. Results from controlled laboratory studies are benchmarked against field experiments and field weathered particles, where data is available. Finally, best practices are summarized and recommendations for experimental design and quantification methods for assessing UV-degradation of plastic in a harmonized matter are provided. Although the primary focus is on the marine environment, the approaches recommended are considered readily transferable to other environmental compartments. This study was conducted within the JPI Oceans ANDROMEDA project, aiming to enhance the understanding of the formation and degradation of micro- and nanoplastic in the marine environment to improve their detectability.

## 2. State-of-the-science

The effects of UV irradiation on plastic degradation in the natural environment have been increasingly studied in the last decades (Gewert et al., 2015), with an overall aim of determining the fragmentation and weathering processes plastic items undergo in the marine environment, and the subsequent impacts these processes have on the fate and effects of plastic and the associated particle and chemical degradation products.

Published literature on marine UV-degradation of plastic was identified from literature searches on (i) Web of Science, with the search string “(ALL=(plastic) OR ALL=(microplastic) OR ALL=(polymer)) AND (ALL=(UV) OR ALL=(photo)) AND (ALL=(irradiation) OR ALL=(degradation)) AND ALL=(marine)”, and (ii) Google Scholar using the search term “(plastic OR microplastic) AND (UV OR photo) AND (irradiation OR degradation) AND marine”. A date limit from 1970 to February 2023, when the literature search was performed, was placed on the Google Scholar search to reflect the date restrictions on Web of Science. A total of 280 and 53,500 results were produced from Web of Science and Google Scholar, respectively. For feasibility, only the first 1000 papers, sorted by relevance, from Google Scholar were taken forward and used in the review process. Thus, a total of 1280 original papers were considered based on their meta-data and abstract.

Review articles, conference proceedings, and non-peer reviewed papers were first excluded and duplicate hits removed. For the remaining papers, inclusion criteria were as follows: (i) UV-degradation studies carried out in a seawater relevant medium, (ii) conventional thermoplastics were included among the studied materials. Both papers on macroplastic and microplastic degradation were included. No eligible studies were identified before 1989. For the purpose of evaluating UV-exposure designs, a total of 63 studies were finally selected and reviewed, representing the period from 1989 to early 2023, with the majority of publications from 2017 onwards. Additionally, eight toxicity studies were reviewed to provide insight into the current knowledge on impacts of UV-exposure on toxicity of resulting chemical leachates.

Fig. 2 provides a general overview of the topics covered in the selected studies. ‘Degradation’ studies include those which qualified or quantified degradation through (i) surface changes, using spectroscopy (e.g., Fourier-Transform infrared; FTIR) and microscopy (e.g., scanning electron microscopy; SEM) techniques, or (ii) mechanical changes, such as tensile strength tests. Studies that monitored the loss of chemical compounds from the plastic through UV irradiation were classified as ‘Leaching’ studies. These included studies that monitored leaching through the release of dissolved organic carbon (DOC), as well as studies that characterized or identified leached compounds, predominantly through targeted analysis approaches using liquid- or gas-chromatography mass spectrometry. Studies investigating in vivo or in vitro toxicity of UV exposed plastics and/or their leachates were
<table><caption>**Table 1**

Summary of the aspects investigated and the methods used in studies dealing with UV weathering of plastics in seawater conditions. Exposure length corresponds to actual exposure time, not simulated/accelerated exposure time.</caption><thead><tr><th>Reference</th><th>Polymer types</th><th>Source of material</th><th>UV source</th><th>Exposure medium</th><th>Exposure length</th><th>Methods to study/quantify fragmentation, surface changes or spectral changes</th><th>Leaching studied</th><th>Other characterization methods</th><th>Comments</th></tr></thead><tbody><tr><td>Andrady (1990)</td><td>LDPE</td><td>Films</td><td>Sunlight</td><td>NSW</td><td>1 y.</td><td>–</td><td>–</td><td>Tensile strength</td><td>–</td></tr><tr><td>Andrady et al. (1993)</td><td>LDPE</td><td>Films</td><td>Sunlight</td><td>NSW (in situ)</td><td>1 w.</td><td>–</td><td>–</td><td>Tensile strength</td><td>Comparison with exposure in freshwater</td></tr><tr><td>Andrady et al. (2022)</td><td>LDPE</td><td>Injection molded LDPE</td><td>UV-B</td><td>ASW</td><td>10 m.</td><td>ATR-FTIR</td><td>–</td><td>Tensile strength</td><td>Comparison with naturally weathered field samples</td></tr><tr><td>Arias-Villamizar and Vázquez-Morillas (2018)</td><td>HDPE, oxodegradable HDPE</td><td>Plastic bags</td><td>Sunlight</td><td>NSW (in situ)</td><td>6 m.</td><td>Visual evaluation, digital images (camera), FTIR</td><td>–</td><td>Tensile strength</td><td>Biofilm formation observed</td></tr><tr><td>Barhoumi et al. (2023)</td><td>Extruded PS</td><td>Environmental sample cut into microplastics (~2-5 mm)</td><td>Sunlight, UV-C (light: dark cycle)</td><td>NSW</td><td>6 m.</td><td>SEM, ATR-FTIR, XRD</td><td>Target analysis of brominated flame retardants (LC-MS/MS)</td><td>–</td><td>Investigated effect of pH, salinity, temperature, and DOM on the leaching of additives<br>Simulated different environmental compartments of the water column (euphotic vs. aphotic zones)</td></tr><tr><td>Beltrán-Sanahuja et al. (2020)</td><td>PLA, PA/PP-EVOH, PE/PET, laminate paper/PLA</td><td>Commercial plastics</td><td>Artificial</td><td>NSW</td><td>1 y.</td><td>ATR-FTIR, weight loss</td><td>–</td><td>DSC</td><td>–</td></tr><tr><td>Biale et al. (2022)</td><td>LDPE, HDPE, PP, PS</td><td>Naturally aged microplastics from beach</td><td>Artificial</td><td>ASW</td><td>2 w.</td><td>–</td><td>Non-target GC-MS, identification of >60 compounds</td><td>–</td><td>–</td></tr><tr><td>Biber et al. (2019)</td><td>PE, PS, PET, Biothene®</td><td>Plastic bags, film</td><td>Sunlight</td><td>NSW (in situ)</td><td>10 m.</td><td>FTIR, weight loss, thickness</td><td>–</td><td>Tensile strength</td><td>–</td></tr><tr><td>Birnstiel et al. (2022)</td><td>LDPE, PS, PLA</td><td>Virgin pellets</td><td>UV-A, UV-B</td><td>ASW</td><td>6 d.</td><td>–</td><td>Fluorescent dissolved organic matter</td><td>Flow cytometry</td><td>–</td></tr><tr><td>Bonifazi et al. (2023)</td><td>PP, HDPE, PS</td><td>Packaging products</td><td>Artificial</td><td>NSW</td><td>~3 m.</td><td>Digital images (camera), stereomicroscope, hyperspectral imaging</td><td>–</td><td>Characterization of seawater prior experiment (pH, conductivity), colorimetric analysis</td><td>Scope: assess recyclability of weathered plastics</td></tr><tr><td>Brandon et al. (2016)</td><td>PET, HDPE, PVC, LDPE, PP, PS</td><td>Plastic pellets</td><td>Sunlight</td><td>NSW</td><td>3 y.</td><td>FTIR</td><td>–</td><td>Crystallinity</td><td>Temperature was monitored</td></tr><tr><td>Cai et al. (2018)</td><td>PE, PP, PS</td><td>Virgin plastic pellets</td><td>UV-A</td><td>ASW</td><td>3 m.</td><td>FTIR, Raman, SEM</td><td>–</td><td>–</td><td>Comparison with exposure in UPW and air</td></tr><tr><td>Chen et al. (2019)</td><td>PVC</td><td>Microplastics (10–300 µm)</td><td>UV-A</td><td>ASW</td><td>~2.5 d.</td><td>SEM-EDS, DLS, ATR-FTIR</td><td>–</td><td>–</td><td>Impacts of salinity and DOM were also investigated</td></tr><tr><td>Chen et al. (2023)</td><td>Mainly PP</td><td>Face masks</td><td>UV-A (light: dark cycle)</td><td>ASW</td><td>~2 w.</td><td>SEM, FTIR, weight loss, Laser particle analyzer (MP size distribution)</td><td>Target analysis of organotin compounds (GC-MS)</td><td>–</td><td>–</td></tr><tr><td>Conradie et al. (2022)</td><td>PP, PET</td><td>Sheets</td><td>UV-A, UV-B</td><td>NSW</td><td>6 w.</td><td>FTIR, SEM, weight loss</td><td>–</td><td>Hydroxyl radicals</td><td>Artificial seawater, 8:16 h light/heat-dark cycle</td></tr><tr><td>Da Costa et al. (2018)</td><td>PE</td><td>Virgin pellets</td><td>UV-C</td><td>ASW</td><td>8 w.</td><td>FTIR, Raman, SEM, weight loss</td><td>–</td><td>TGA</td><td>Pre-treatment with UV before seawater exposure to simulate beach debris</td></tr><tr><td>De-la-Torre et al. (2022)</td><td>PP, LDPE</td><td>Personal protective equipment</td><td>Sunlight</td><td>NSW</td><td>2 m.</td><td>FTIR, XRD, SEM-EDS</td><td>–</td><td>–</td><td>Comparison with exposure in deionized water</td></tr><tr><td>Dhavamani et al. (2022)</td><td>HDPE, LDPE, recycled PE</td><td>Film grade polymers</td><td>UV-A</td><td>ASW</td><td>5 d.</td><td>–</td><td>Target analysis of phthalates (GC-MS)</td><td>FTIR prior to the experiment</td><td>Comparison with exposure in distilled water<br>Leaching and adsorption of phthalates from PE was investigated</td></tr></tbody></table>

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<table><caption>Table 1 (continued)</caption><thead><tr><th>Reference</th><th>Polymer types</th><th>Source of material</th><th>UV source</th><th>Exposure medium</th><th>Exposure length</th><th>Methods to study/quantify fragmentation, surface changes or spectral changes</th><th>Leaching studied</th><th>Other characterization methods</th><th>Comments</th></tr></thead><tbody><tr><td>Fernández-González et al. (2021)</td><td>HDPE, LDPE, PET, PP, PS</td><td>Virgin powders (~300 µm) & pellets (~3 mm)</td><td>Artificial</td><td>NSW</td><td>10 w.</td><td>FTIR, SEM</td><td>–</td><td>–</td><td>Comparison with weathering under dry conditions</td></tr><tr><td>Gerritse et al. (2020)</td><td>PE, PS, PP, Latex, PET, PU, cellulose acetate</td><td>Consumer products (bags, trays, cups, cigarette butts)</td><td>Visible, UV-B (light:dark cycle)</td><td>ASW</td><td>~1 y.</td><td>Weight loss</td><td>–</td><td>Electrical resistance measurements, characterization of microbial biofilm</td><td>–</td></tr><tr><td>Han et al. (2021)</td><td>PE, PP, PET</td><td>Virgin pellets (3 size fractions)</td><td>UV-A</td><td>ASW</td><td>1.7 h</td><td>ATR-FTIR</td><td>–</td><td>–</td><td>Cu(II) adsorption experiments were conducted</td></tr><tr><td>Hingant et al. (2023)</td><td>PP</td><td>Oyster spat collectors</td><td>UV-A, UV-B</td><td>ASW</td><td>5 m.</td><td>ESEM, FTIR</td><td>–</td><td>DSC</td><td>Comparison with in situ weathering</td></tr><tr><td>Iniguez et al. (2018)</td><td>PE, PP, PET, Nylon</td><td>Consumer goods</td><td>UV-C</td><td>NSW</td><td>6.5 m.</td><td>AFM, SEM</td><td>–</td><td>Tensile strength, TGA</td><td>Scope: assess recyclability of weathered plastics</td></tr><tr><td>Kalogerakis et al. (2017)</td><td>HDPE</td><td>Plastic bags</td><td>Sunlight</td><td>NSW</td><td>6 m.</td><td>Molecular weight measurements, FTIR, weight loss, image processing of photographs</td><td>–</td><td>Tensile strength</td><td>Temperature and sunlight intensity were monitored</td></tr><tr><td>Khoironi et al. (2020)</td><td>PP</td><td>Plastic cups</td><td>Sunlight</td><td>NSW (in situ)</td><td>~3 m.</td><td>FTIR, SEM, EDS</td><td>–</td><td>–</td><td>–</td></tr><tr><td>La Mantia et al. (2022)</td><td>PP</td><td>Virgin polymer</td><td>UV-B</td><td>NSW</td><td>~2.5 w.</td><td>UV-vis absorption spectra, ATR-FTIR</td><td>–</td><td>Tensile strength, elongation, DSC, oxygen transmission rate, molecular weight</td><td>Comparison with exposure in distilled water and air</td></tr><tr><td>Lambert et al. (2013)</td><td>Latex</td><td>Films</td><td>Sunlight</td><td>ASW</td><td>~7 m.</td><td>ATR-FTIR, NTA, weight loss</td><td>DOC</td><td>pH</td><td>Comparison with exposure in artificial freshwater media</td></tr><tr><td>Leonas (1993)</td><td>LDPE, PS</td><td>Films</td><td>UV-A</td><td>ASW</td><td>10 d.</td><td>–</td><td>–</td><td>Tensile strength, elongation</td><td>Comparison with exposure in freshwater and air</td></tr><tr><td>Mao et al. (2020)</td><td>PS</td><td>Pristine spheres (0.1 µm)</td><td>UV-A</td><td>ASW</td><td>3 m.</td><td>SEM, FTIR, X-ray photoelectron spectroscopy</td><td>–</td><td>Effect of aging on adsorption of heavy metals by PS</td><td>Comparison with exposure in pure water and air</td></tr><tr><td>Naik et al. (2020)</td><td>HDPE, Nylon, PP, high impact PS</td><td>Virgin pellets, environmental samples for comparison</td><td>UV-C</td><td>ASW</td><td>5 d.</td><td>ATR-FTIR, SEM, Raman</td><td>–</td><td>–</td><td>Microplastic particle and fiber generation were investigated</td></tr><tr><td>Nelson et al. (2021)</td><td>PE, PET</td><td>Consumer products</td><td>Sunlight</td><td>NSW</td><td>6 w.</td><td>–</td><td>–</td><td>Seawater characterization (pH, conductivity), total solar irradiance, plastic film transmittance, biomass formation</td><td>–</td></tr><tr><td>Niaounakis et al. (2019)</td><td>PE, PET, PAPE, Mater-Bi®</td><td>Plastic bags, film</td><td>Visible, UV-A</td><td>NSW</td><td>5 d.</td><td>FTIR</td><td>–</td><td>Tensile strength, DSC</td><td>Comparison with prolonged hydrolysis (without UV irradiation)</td></tr><tr><td>O'Brine and Thompson (2010)</td><td>PE, biobased PE</td><td>Plastic bags</td><td>Sunlight</td><td>NSW (in situ)</td><td>40 w.</td><td>Light microscopy</td><td>–</td><td>Tensile strength, UV light transmittance</td><td>Formation of biofilm was investigated</td></tr><tr><td>Paluselli et al. (2019)</td><td>PE, PVC</td><td>Black garbage bag, cable insulation</td><td>Artificial</td><td>NSW</td><td>3 m.</td><td>SEM</td><td>Target analysis of phthalates (GC-MS), DOC</td><td>–</td><td>Biotic versus abiotic conditions were tested</td></tr><tr><td>Pegram and Andrady (1989)</td><td>LDPE, PP, Latex</td><td>Commercial suppliers, fishing net</td><td>Sunlight</td><td>NSW (in situ)</td><td>1 y.</td><td>–</td><td>–</td><td>Tensile strength</td><td>Comparison with exposure in air</td></tr></tbody></table>

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<table><caption>Table 1 (continued)</caption><thead><tr><th>Reference</th><th>Polymer types</th><th>Source of material</th><th>UV source</th><th>Exposure medium</th><th>Exposure length</th><th>Methods to study/quantify fragmentation, surface changes or spectral changes</th><th>Leaching studied</th><th>Other characterization methods</th><th>Comments</th></tr></thead><tbody><tr><td>Quade et al. (2022)</td><td>PE, two biobased materials</td><td>Plastic bags</td><td>Sunlight</td><td>NSW</td><td>4 m.</td><td>Visual and tactile evaluation, weight loss</td><td>Not analyzed, but additive release was suspected to be responsible for observed toxicity</td><td>Water and sediment characterization (NH<sub>4</sub><sup>+</sup>, NO<sub>2</sub><sup>-</sup>, NO<sub>3</sub><sup>-</sup>, PO<sub>4</sub><sup>3-</sup>, TOC)</td><td>-</td></tr><tr><td>Rani et al. (2017b)</td><td>EPS</td><td>Spherules detached from new EPS buoys, varying sizes</td><td>Sunlight</td><td>NSW (in situ)</td><td>6 m.</td><td>-</td><td>Target analysis (LC-MS/MS)</td><td>Macroscopic observation (biofouling)</td><td>-</td></tr><tr><td>Ranjan and Goel (2019)</td><td>LDPE</td><td>Virgin packaging sheet</td><td>Artificial</td><td>ASW</td><td>3 m.</td><td>FTIR, SEM, EDS, AFM</td><td>-</td><td>Tensile strength, elongation, Young's modulus</td><td>-</td></tr><tr><td>Ranjan and Goel (2021)</td><td>PP</td><td>New plastic sheets</td><td>UV-A</td><td>ASW</td><td>1 m.</td><td>AFM, SEM, FTIR</td><td>-</td><td>Tensile strength, elemental analysis, TGA</td><td>Scope: assess recyclability of weathered plastics</td></tr><tr><td>Romera-Castillo et al. (2022)</td><td>LDPE</td><td>Virgin film, mix of aged plastics from environment (65 % PE, 30 % PP, 5 % PVC)</td><td>UV-A, UV-B</td><td>NSW</td><td>6 d.</td><td>-</td><td>DOC</td><td>FTIR & SEM to characterize plastics from the environment prior to the experiment, bacterial abundance (flow cytometry)</td><td>-</td></tr><tr><td>Rummel et al. (2019)</td><td>PE, PET, PS, PP</td><td>Virgin polymers, samples from electronic equipment</td><td>UV-A, UV-B</td><td>ASW</td><td>4 d.</td><td>-</td><td>Target analysis of mono- & dicarboxylic acids (LC-MS)</td><td>-</td><td>Cell-based bioassays were conducted</td></tr><tr><td>Rummel et al. (2022)</td><td>PE, PET, PS, PP</td><td>Virgin polymers, samples from electronic equipment</td><td>UV-A, UV-B</td><td>ASW</td><td>4 d.</td><td>-</td><td>Target analysis of mono- & dicarboxylic acids (LC-MS)</td><td>-</td><td>Cell-based bioassays were conducted</td></tr><tr><td>Sait et al. (2021)</td><td>PET, PA, PAN</td><td>Microfibers from commercial textile producers</td><td>Artificial (daylight filter)</td><td>NSW</td><td>10 m.</td><td>Light microscopy, FTIR, SEM</td><td>Non-target (GC-MS), target analysis of benzothiazoles and benzophenones (UPLC-MS/MS), ICP-MS</td><td>-</td><td>Comparison with exposure in artificial freshwater</td></tr><tr><td>Sarno et al. (2021)</td><td>PET fibers</td><td>Fleece material</td><td>Artificial (daylight filter)</td><td>NSW</td><td>1 m.</td><td>SEM</td><td>Target analysis of terephthalic acid and ethylene glycol (LC-MS/MS)</td><td>-</td><td>Aimed to produce microplastic fiber reference materials, comparison to accelerated hydrolysis</td></tr><tr><td>Shashoua et al. (2022)</td><td>PUR</td><td>Household sponges</td><td>Sunlight</td><td>NSW (in situ)</td><td>2 y.</td><td>FTIR, weight loss</td><td>No leaching but extraction of degradation products and plastic additives from weathered sponges (GC-MS)</td><td>Flexibility</td><td>Investigated effect of different depths</td></tr><tr><td>Sorasan et al. (2022)</td><td>PE, PP, PS</td><td>Environmental samples (1–5 mm)</td><td>Visible, UV-A, UV-B</td><td>ASW</td><td>15 d.</td><td>ATR-FTIR, particle size distribution</td><td>-</td><td>Flow cytometry, TOC, DSC, TGA</td><td>Nanoplastic generation was investigated</td></tr><tr><td>Sørensen et al. (2021)</td><td>PET, PA</td><td>Microfibers from commercial textile producers</td><td>Artificial (daylight filter)</td><td>NSW</td><td>~ 2 m.</td><td>Light microscopy, SEM</td><td>Non-target and target (GCMS) for various additives and degradation products</td><td>-</td><td>-</td></tr><tr><td>Takahashi et al. (2022)</td><td>PET</td><td>Scraps from postconsumer bottles</td><td>UV-B</td><td>ASW</td><td>~3.5 d.</td><td>-</td><td>Terephthalate and related compounds, DOC</td><td>-</td><td>-</td></tr><tr><td>Tang et al. (2018)</td><td>PVC</td><td>Virgin pellets (~100 µm)</td><td>UV-B, sunlight</td><td>ASW</td><td>7 m.</td><td>SEM-EDS, ATR-FTIR</td><td>-</td><td>Surface area, pore volume and average pore size were determined using nitrogen adsorption-desorption</td><td>-</td></tr><tr><td>Tang et al. (2019)</td><td>PP</td><td>Virgin pellets (~4 mm)</td><td>UV-B, sunlight</td><td>ASW</td><td>1.5 y.</td><td>ESEM, ATR-FTIR</td><td>-</td><td>-</td><td>Photo- and thermal-oxidative degradation were investigated</td></tr></tbody></table>

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<table><caption>Table 1 (continued)</caption><thead><tr><th>Reference</th><th>Polymer types</th><th>Source of material</th><th>UV source</th><th>Exposure medium</th><th>Exposure length</th><th>Methods to study/quantify fragmentation, surface changes or spectral changes</th><th>Leaching studied</th><th>Other characterization methods</th><th>Comments</th></tr></thead><tbody><tr><td>Tang et al. (2022)</td><td>POM</td><td>Virgin pellets</td><td>Sunlight</td><td>ASW</td><td>~1 y.</td><td>SEM, FTIR, weight loss</td><td>–</td><td>–</td><td>–</td></tr><tr><td>Tetu et al. (2019)</td><td>HDPE, PVC</td><td>Grocery bag, plastic matting</td><td>Artificial</td><td>ASW</td><td>5 d.</td><td>–</td><td>ICP-OES, ICP-MS, non-target LC-MS/M</td><td>–</td><td>Toxicity of leachates was tested on photosynthetic bacteria</td></tr><tr><td>Tong et al. (2022)</td><td>PBAT, PBS, PLA, PE, PS, PVC</td><td>Pellets (2–3 mm), plastic bags</td><td>UV-A (light: dark cycle)</td><td>ASW</td><td>3 w.</td><td>SEM, fluorescence microscope</td><td>–</td><td>–</td><td>Toxicity of particles was tested on copepods</td></tr><tr><td>Tosin et al. (2012)</td><td>LDPE, Mater-Bi®</td><td>Carrier bags</td><td>Sunlight</td><td>NSW</td><td>2 y.</td><td>–</td><td>–</td><td>Tensile strength, elongation at break</td><td>–</td></tr><tr><td>Wang et al. (2020)</td><td>PS</td><td>Pristine fragments</td><td>UV-A, UV-B</td><td>NSW</td><td>2 m.</td><td>SEM, FTIR</td><td>GC-MS</td><td>–</td><td>Leachates and microplastics were used to test toxicity, bioaccumulation was assessed</td></tr><tr><td>Wang et al. (2021)</td><td>PS, PVC</td><td>Virgin polymers</td><td>UV-A</td><td>NSW</td><td>3 m.</td><td>SEM, FTIR</td><td>–</td><td>Zeta potential</td><td>Combined toxicity of microplastics and heavy metals was investigated</td></tr><tr><td>Welden and Cowie (2017)</td><td>PP, PE, Nylon</td><td>Ropes</td><td>Sunlight</td><td>NSW (in situ)</td><td>1 y.</td><td>–</td><td>–</td><td>Tensile strength, elongation, weight</td><td>Study conducted at 10 m depth (benthic conditions), water temperature and light levels monitored</td></tr><tr><td>Wu et al. (2020)</td><td>PS</td><td>Virgin pellets (~4 mm)</td><td>UV-A</td><td>NSW</td><td>3 m.</td><td>SEM, FTIR, x-ray diffraction</td><td>–</td><td>Sorption of BDE-47</td><td>–</td></tr><tr><td>Wu et al. (2021a)</td><td>PP</td><td>Food packaging</td><td>UV-A (light: dark cycle)</td><td>ASW</td><td>12 d.</td><td>SEM, FTIR, x-ray diffraction</td><td>No leaching, but additive and NIAS content in particles: non-target (GC-MS), target (GC-MS)</td><td>TOC, TG-DSC, hydroxyl radicals (HPLC)</td><td>Additional tests to investigate influence of antioxidant on microplastic fragmentation</td></tr><tr><td>Wu et al. (2021b)</td><td>PP</td><td>Pristine microplastics</td><td>UV-A</td><td>NSW</td><td>~3 w.</td><td>SEM, FTIR</td><td>–</td><td>TG-DSC, formation of reactive oxygen species</td><td>Inhibitory effect of Cl<sup>-</sup> for PP aging in seawater was investigated</td></tr><tr><td>Wu et al. (2022)</td><td>PP, PS, thermoplastic PUR</td><td>Virgin microplastic spheres</td><td>Artificial (daylight filter)</td><td>ASW</td><td>2 m.</td><td>–</td><td>DOC, chromophoric dissolved organic matter, dissolved nitrogen</td><td>–</td><td>Effects of different spectral compositions of the incident light were tested</td></tr><tr><td>Zhao et al. (2022)</td><td>PE, PP, biodegradable & non-biodegradable blends</td><td>Mechanical grinding of commercially available polymers & polymer blends</td><td>Artificial (daylight filter)</td><td>NSW</td><td>2 m.</td><td>SEM, FTIR, confocal laser scanning microscopy (size distribution), molecular weight</td><td>–</td><td>Contact angle, zeta potential</td><td>–</td></tr><tr><td>Zhu et al. (2020)</td><td>PE, PP, EPS</td><td>Postconsumer microplastics from the North Pacific Gyre</td><td>UV-A, UV-B</td><td>NSW</td><td>~3 m.</td><td>SEM, FTIR, weight loss, optical photographs</td><td>DOC</td><td>Flow cytometry</td><td>Consumption of released DOC by marine bacteria was observed</td></tr></tbody></table>

NSW = natural seawater, ASW = artificial seawater, SEM = scanning electron microscopy, ESEM = environmental scanning electron microscopy, AFM = atomic force microscope, DLS = dynamic light scattering, TOC = total organic carbon, DOC = dissolved organic carbon, DOM = dissolved organic matter, TGA = thermogravimetric analysis, TG-DSC = thermo gravimetry-differential scanning calorimetry, POM = polyoxymethylene, EVOH = ethylene-vinyl-alcohol, EPS = expanded polystyrene, EDS = energy dispersion X-ray spectrometer, XRD = X-ray diffraction, PAPE = polyamide-polyethylene, XPS = X-ray photoelectron spectroscopy, NTA = nanoparticle tracking analysis, NIAS = non-intentionally added substances, BDE-47 = 2,20,4,40-tetrabromodiphenyl ether, d. = days, w. = weeks, m. = months, y. = years.
<figure><img src="image_3.png" /><figcaption>Fig. 1. Exposure of plastics to UV radiation influences the magnitude of chemical leaching as well as the chemical composition of the leachate.</figcaption></figure>

<figure><img src="image_4.png" /><figcaption>Fig. 2. Number of UV exposure studies of plastics in seawater per year and count of the type of research in each year. Total number of individual studies = 63, total number of category allocations = 73. *until February 2023.</figcaption></figure>

classified as ‘Toxicity’ studies. Finally, studies investigating the adsorption behavior of UV-exposed plastic materials in seawater were categorized as ‘Adsorption’ studies. Of the 63 papers included in this review, 51 were categorized as ‘Degradation’ studies, 21 as ‘Leaching’ studies, 7 as ‘Toxicity’ and 5 as ‘Adsorption’ studies. Where an individual paper contained data relevant to more than one category, the paper was assigned to all relevant categories. In total, 10 papers were placed into two of the aforementioned categories, most of which were published in the last five years, with the exception of Lambert et al. (2013). Since 2017, the number of studies on UV degradation of plastics under marine conditions has increased, from 0 to 2 studies per year to 16 in 2022 (Fig. 2). This increase is expected to continue, given that within the first two months of 2023 the number of published studies already represented one quarter of the total number of studies published in 2022 on this subject. This highlights the need for rapid harmonization of protocols and methods.

### 3. Experimental design of UV-degradation studies

An understanding of UV-degradation of plastic is critical for testing and improving material stability (e.g., by adding UV stabilizers or coatings) within the plastic industry. As a result, several ISO and ASTM

standards have been developed for assessing the resistance of plastic/ polymer materials to environmental UV-degradation (e.g., ASTM D4329-21, ASTM D5272-08(2021), ISO 4892-1:2024, ISO 10640:2011). However, there are no standard methods specifically for designing and conducting studies that represent UV-degradation processes in the marine environment where there is a focus on quantifying particulate and chemical degradation product emissions from plastic. Despite this situation, read-across and lessons from material performance testing using standard methods could and should be implemented in the design of environmental fate studies. At the same time, end-points of analysis and analysis conditions should ideally be selected to be representative of actual marine conditions.

When comparing the exposure approaches used for UV-degradation assessment in the 63 studies reviewed herein, it was found that each study had a unique set-up. This leads to difficulties in meaningfully comparing data generated from different studies and highlights the need for standardization of marine environmental UV-degradation assessment methods. Across the 63 studies reviewed, 41 used an artificial UV source (65 %), 18 exposed samples to natural sunlight (29 %), and four studies used both artificial and natural UV sources (6 %) (Fig. 3). Importantly, the exposure conditions employed varied widely across studies, including the use of filters to simulate natural sunlight or
<figure><img src="image_5.png" /><figcaption>**Fig. 3.** (A) shows the number of studies using natural, artificial, or natural and artificial UV-sources for plastic exposure experiments, (B) shows the proportion of studies using an artificial seawater or natural seawater medium and (C) shows the most common polymer types used in the studies. The ‘Other’ category includes polymers such as polyamide (PA), polyurethane (PUR), polycarbonate (PC), biopolymers polybutylene succinate (PBS), polyhydroxybutyrate (PHB) and poly (butylene adipate-co-terephthalate) (PBAT) and tyre rubbers (see Table 1 for the full list of polymers used in each study).</figcaption></figure>

sequential exposure to simulate a natural 24-h light cycle.

Among the studies using both artificial and natural UV sources, Wu *et al.* (2021a, 2021b) used a natural UV exposure to confirm that 12 days of accelerated weathering under their laboratory conditions corresponded to approximately 90 days of natural outdoor weathering. This is a good example of how the environmental relevance of laboratory exposure conditions can be tested, at least in a simplified matter. Barhoumi *et al.* (2023) could discern an increased hydrophobicity of naturally weathered samples compared to samples subjected to artificial UV exposure, though this is likely due to the biofilm which developed on these particles, and which is connected mainly to them having been suspended in natural marine waters. Importantly, Tang *et al.* (2019) observed that cracks formed inside PP pellets subjected to sunlight irradiation, while under artificial UV irradiation cracks were not found in the interior of the particles. This suggests that artificial UV exposure does not accurately simulate the effects that natural sunlight irradiation has on plastic particles.

A total of 9 studies (14 %) carried out experiments in situ open seawater systems, where plastics were deployed directly into the marine environment and subsequently exposed to both natural seawater and sunlight. Across the 9 in situ degradation studies, the various test materials were deployed at different depths, ranging from floating at the sea surface to a depth of 10 m, therefore significantly impacting the level of UV solar irradiation exposure the test materials received. Laboratory-based UV-degradation studies were primarily conducted in closed seawater systems rarely exceeding 100 mL that were exposed using solar/UV simulation chambers. In such exposures, the plastic material and seawater is usually placed in quartz tubes, which allow UV irradiation to pass through. Irradiation intensity, exposure time and chamber temperature can be specified and monitored. Optical filters are typically used to limit the emitted irradiation to a specific wavelength range, such as simulating the natural sunlight spectrum (e.g. UVA and UVB; wavelengths >280 nm). The small sample volumes reflect the limited exposure area available in most lab-scale solar simulation systems and the need to include multiple materials, time points, blanks/controls in a single study conducted over long periods of time (weeks-months).

An artificial seawater medium was used in 28 (44 %) studies, while 35 (56 %) of studies used natural seawater (either in the laboratory or in the field). Most studies using natural seawater in a laboratory setting pre-processed the seawater by filtration (using various pore sizes) and/or sterilization (pre-treatment with UV-C light or addition of HgCl₂). Depending on material properties (e.g. polymer type), the time scales required to achieve significant degradation can range from months to years (Sait *et al.*, 2021). Approximately half (52 %) of studies conducted exposures over a duration of months (1–11), while the remaining

studies were conducted over a period of days (1–7), weeks (1–40) or years (1–2). As it is usually impractical to run long-term studies over several years, accelerating exposures may be desirable. The process can be sped up by conducting the UV exposure continuously and by using high UV intensities, although kept within the wavelength range of natural sunlight. Even under accelerated conditions, it is usually necessary to conduct degradation experiments over longer time periods to induce observable changes in most thermoplastics. Shorter wavelengths are more effective at degrading plastic materials and exposure to light containing UVC wavelengths (200–280 nm) can significantly speed up the degradation process (Doğan, 2021; Lee *et al.*, 2020a, 2020b). However, it is important to note that UVC exposure may lead to the formation of different degradation products and oxygenated radicals compared to exposures conducted within the natural sunlight wavelength range (Doğan, 2021), thus presenting with lower environmental relevance compared to simulated daylight exposures.

Individual studies focused on plastic chemical leaching may not attempt to qualify or quantify degradation of the plastic material itself, but the use of UV to facilitate plastic chemical leaching would lead to photodegradation of the plastic (e.g., Dhavamani *et al.*, 2022). Similarly, adsorption studies often characterize the degree of plastic degradation as this can affect adsorption to the material (e.g., Mao *et al.*, 2020).

The different methodological approaches have their individual strengths and weaknesses, and the appropriate choice often depends on the objective of the study. For example, using natural sunlight and natural marine waters will allow the best reproduction of environmental weathering conditions, while simplifying the process by using controlled exposure conditions such as artificial seawater will allow to better discern the impact of individual factors on UV degradation of plastics. For the comparability of studies, a certain level of harmonization or standardization in laboratory studies is nonetheless recommended. One approach could be for each study to include a control sample using standardized plastic pellets and standardized artificial seawater, which would allow to discern the impact the basic exposure conditions (temperature, irradiation, duration) from each study have on the plastics. Furthermore, the inclusion of both artificial UV and natural sunlight exposure in studies is highly recommended.

## 4. Selection of plastic test materials impacts outcomes

The physicochemical properties of the test materials chosen for use in UV-degradation experiments can also have a strong influence on the outcomes of a study. The plastic materials investigated across the 63 studies represented a range of material types, from virgin polymer films to environmentally collected macro- and microplastics. A majority of
studies (62 %) primarily investigated macroplastic, while the remainder employed microplastics (< 5 mm diameter) as the start material (either in the form of virgin pellets or cryomilled/otherwise prepared/degraded particles). Particle or material shape/morphology and composition formulation (polymer type and additive chemical profile) can play important roles in UV-degradation behavior. For example, films/sheets may be more prone to damage by UV than granulates due to their increased surface to volume ratio. Similarly, assessing UV-degradation with virgin polymers (i.e., without additives) is likely to have low environmental relevance due to the lack of stabilizing chemicals that are found in most plastic consumer products.

The most used plastics across the studies were polyethylene (PE; 65 %), polypropylene (PP; 43 %), polystyrene (PS; 30 %), polyethylene terephthalate (PET; 24 %), polyvinyl chloride (PVC; 10 %), polylactic acid (PLA; 6 %) and nylon (5 %) (Fig. 2). Many studies used multiple plastic types, with some comparing virgin products to post-consumer products or environmental samples. The plastic materials used in each study were often justified by referencing the predominant use and high production volume of a certain polymer type or its abundance in the marine environment. Further justification was used in studies where new materials were compared to environmental samples. For example, Romera-Castillo et al. (2022) used new LDPE as a comparison along with particles collected in the environment, because LDPE represented 65 % of the polymers detected in the environmental samples collected for the study.

Polymer selection should be carefully considered, particularly with respect to the intrinsic physicochemical properties and the relevance of certain materials with respect to their behavior in the marine environment. For example, PVC was used in six studies, but its higher density than seawater means that it will generally sink to the seafloor and is therefore less exposed to UV irradiation in the natural marine (and aquatic in general) environment. This may however not be the case for inter-tidal (beach) environments, where UV exposure is relevant for all polymer type materials present. Studies of denser polymers such as PVC could thus be encouraged to use a methodology similar to Conradie et al. (2022), where particles were pre-treated with UV to simulate exposure on land, before further degradation studies were conducted in seawater. This approach may more accurately reflect the movement of PVC plastics through the environment. A similar approach could be used for styrene-butadiene rubber (SBR), an indicator of tire wear particles that currently remains understudied regarding its degradation potential and pathway in the marine environment.

While four (6.3 %) of the studies utilized partially degraded test materials collected from the environment in their UV-degradation studies, the unknown provenance and history can make interpreting results challenging.

## 5. Degradation: Changes in surface chemistry and morphology

The primary changes in surface morphology and integrity of plastic exposed to UV irradiation occur at the material surface, typically starting with chemical degradation of the polymer (and photosensitive additive chemicals) (Hu et al., 2025). UV exposure in the presence of oxygen can result in UV catalyzed oxidation (photooxidation) of the polymers, leading to new functional groups being formed. Photooxidation is initiated when a photon is absorbed by a chemical bond, most likely a weaker C—H bond, which generates polymer chain scission or free radical creation. Oxygen then reacts with the formed radicals to form peroxy radicals. This addition withdraws hydrogens from the vicinal chains, resulting in the formation of hydroperoxides and more free radicals (Alimi et al., 2022; Min et al., 2020). It has been argued that conditions within the upper layers of the oceans limit photooxidative fragmentation, in contrast to beach environments (Andrady, 2022).

Most studies use spectroscopy (FTIR or Raman) to observe and quantify the increase of oxygen-containing functional groups on the surface of plastic materials by monitoring the formation of C—O and

C=O bonds (Lin et al., 2020; Liu et al., 2019; Miranda et al., 2021; Moreira et al., 2021; Ren et al., 2021; Sait et al., 2021). The carbonyl index (CI) is the most commonly used indicator to measure the chemical oxidation of polyolefins such as PE and PP (Allen and Edge, 1992; Mellor et al., 1973) and has also been used for many other polymer types. CI is determined by normalizing the height or area of the carbonyl peak to the height or area of a stable reference peak. Hydroxyl indices, calculated in a similar way as CI, have also seen increased use in quantifying oxidation of polymer following UV exposure (Conradie et al., 2022). Despite large variations in experimental design, most of the 63 studies observed changes in the IR spectral peaks for at least some of the plastics tested, reflecting photooxidation and shifts in the carbonyl indices (CIs). For example, Sait et al. (2021) exposed differently colored PET, nylon (polyamide; PA) and polyacrylic (PAN) fibers in artificial seawater and freshwater to a xenon lamp (1500 W) with a daylight filter (300–400 nm) at an intensity of 65 W/m². Attenuated Total Reflectance-FTIR (ATR-FTIR) showed a decrease in the absorbance at 1711, 1238, and 1090 cm⁻¹ for PET fibers, indicating UV-induced photooxidation. This coincided with a decrease in the CI after 5 months of exposure. Increased absorption at 3275, 2921, and 1631 cm⁻¹ was observed for PA fibers, which was attributed to accelerated UV-degradation causing the formation of imide, accompanied by a temporal increase in CI over 10 months. No measurable shifts in the absorbance nor change in the CI was observed for PAN fibers.

Miranda et al. (2021) investigated the UV degradation of LDPE, PET and unplasticized PVC (uPVC) microplastics. In the first exposure, wet (Milli-Q) and dry microplastics were placed in quartz Petri dishes and exposed for up to 96 h using a UV-C lamp (Heraeus TNN 15/32, 15 W) with an emission at 253.7 nm. The second exposure used a solar simulator (Solarbox 1500e, 1500 W) operated at an intensity of 416 W/m², where dry microplastics in quartz Petri dishes were exposed to 6 h/day of irradiation (for a maximum of 120 h). Finally, an outdoor rooftop experiment was conducted over 13 weeks (spring/summer conditions in Porto, Portugal), where dry and wet (sieved rainwater) microplastics were placed on filters and contained in Petri dishes. UV-C irradiation and the rooftop exposure led to an increase in the CI for both LDPE and uPVC (at the band 1712–1722 cm⁻¹), while the solar simulator increased the CI for LDPE only. ATR-FTIR analysis was unable to detect measurable changes in the oxidation of PET in any of the exposure scenarios, owing to the strong carboxylic acid peak at 1714 cm⁻¹. The above studies show the importance of using approaches such as CI for quantifying surface oxidation when assessing the effects of UV exposure in (micro)plastics.

Several studies have observed UV-induced physical surface changes in plastics, mainly in the form of pitting, cracks and wrinkles (Chen et al., 2020; Fan et al., 2021; Hüffer et al., 2018; Iniguez et al., 2018; Lin et al., 2020; Mao et al., 2020). Sørensen et al. (2021) observed changes in the surface morphology of both synthetic and natural (wool) microfibers when exposed to UV light in natural seawater media. PA and PET fibers exhibited pitting at the surface after 56 days of UV exposure, possibly reflecting regions of the surface characterized as being more amorphous. The wool fibers did not exhibit any pitting, but instead the surface seemed to be smoothed after exposure. Significant fragmentation was only observed for the PET and wool fibers at the end of the 56-day exposure period. Miranda et al. (2021) observed randomly distributed cracks and deformation on the surface of LDPE when exposed to 13 weeks of UV light, while PET was observed to begin flaking. Similarly, Ainali et al. (2021) observed roughness, holes, and defects being formed in LDPE, HDPE, PP, and PS films after 30 and 60 days of UV exposure. These physical changes in the surface layers of plastic materials represent the first steps towards fragmentation but are difficult to use as a basis for quantifying the degree of degradation.

To fully account for surface degradation of plastic or microplastic, both visualization techniques and chemical techniques should be applied. In most cases, the reviewed studies have used a combination of SEM and FTIR for this purpose, and these techniques seem reasonably
widely available. Replacing FTIR with Raman or hyperspectral imaging for evaluation of chemical surface changes may be appropriate, while for larger plastic items other optical techniques may serve to evaluate surface degradation – although with the understanding that they will not necessarily be able to measure or image the initial stages of the degradation process.

## 6. Degradation: UV-induced fragmentation and formation of micro- and nanoplastics

Surface embrittlement caused by UV irradiation is a critical initial step in the fragmentation of plastic litter, where secondary microplastics and nanoplastics are formed via the fracture and mild abrasion of highly UV-degraded plastic surfaces (Andrady et al., 2019; Jahnke et al., 2017). As microplastics and nanoplastics represent an environmental and human health concern due to their small size making them more biologically available than larger plastic (Gigault et al., 2021; Macali et al., 2018; Rochman et al., 2015; Van Cauwenberghe et al., 2015), it is important to understand their formation and release as a result of UV-degradation of plastic. In addition, studies have shown that these micro- and nanoplastics experience quite different deposition, homoaggregation, and hetero-aggregation processes in the marine environment (Alimi et al., 2018).

Owing to limitations with available characterization techniques and sampling methods, only a few studies have investigated the formation of small micro- and nanoplastics in water following UV exposure. Lambert and Wagner (2016) used nanoparticle tracking analysis (NTA) to quantify the number of small particles emitted from PS coffee cup lids when exposed to UV-visible radiation. The results showed significant differences in the number of particles between the PS sample (containing the coffee cup lid) and the control (containing only demineralized water) after just 10 days of UV exposure. However, studies investigating the formation of nanoplastics under UV exposure in marine conditions are largely lacking so far. A study reporting the formation of nanoplastics during exposure of microplastics collected at sea and exposed to UV irradiation in ultrapure water suggests however, that this process also occurs in the natural marine environment (Gigault et al., 2016). A major limitation in the study of nanoplastics is the lack of analytical techniques for detection and quantification of these small carbon-based particles in a carbon-rich environment. Techniques such as optical photothermal IR spectroscopy or pyrolysis-GC-MS are promising and are being increasingly used in environmental studies of nanoplastics.

## 7. Degradation: Changes in bulk plastic properties and integrity upon UV-exposure

UV-degradation has the potential to modify the bulk properties of plastic materials, including crystallinity, glass transition temperature (T~g~) and phase transition temperature (Ainali et al., 2021). Particularly when applying accelerated degradation techniques, it is important to keep track of changes in material properties that may influence the fate of the particles in the environment and ensure non-significant differences between accelerated and natural UV-weathering. A number of studies have investigated UV-degradation of plastics using techniques that analyze the bulk material as opposed to only the surface. Commonly measured parameters include crystallinity (typically analyzed by X-ray diffraction) and thermal properties, analyzed by differential scanning calorimetry (DSC) or thermal gravimetry analysis (TGA). Ainali et al. (2021) showed that UV exposure decreased the melting temperature of the more crystalline HDPE, LDPE, and PP, while a decrease in the glass transition temperature (T~g~) was observed for the more amorphous PS. After 60 days of UV exposure, there was no significant change in the crystallinity of HDPE or LDPE. In contrast, PP began to exhibit a shift from α-type crystals to pseudo-hexagonal β-form crystals after 10 days of UV exposure. Additionally, the degree of PP crystallinity increased over the first 10 days of UV exposure, before decreasing sharply. As with PS,

the PP was found to be amorphous at the start of the experiment, but UV exposure caused a broadening of diffraction peaks that indicated macromolecule chain packing and an increased crystallinity.

Ainali et al. (2021) also studied the mechanisms of plastic UV-degradation by analyzing the ratio between low molecular weight compounds and higher molecular weight hydrocarbons using pyrolysis coupled with gas chromatography and mass spectroscopy (pyGC-MS). The use of pyGC-MS also enabled the identification of oxidation products. Here it is important to note that attempted characterization of surface chemical changes with techniques such as pyGC-MS are not appropriate, as the bulk material is analyzed. Another commonly used method for molecular size characterization is gel-permeation chromatography (GPC). While this technique is widely used within the polymer industry, its potential application for determining changes in molecular size caused by UV-exposure appears to have been largely overlooked in many environmental degradation studies. Although the environmental relevance of changes in plastic bulk material properties may not be obvious, it can lead to a weakening of the material that facilitates fragmentation and further degradation. Changes in bulk material integrity can be measured as tensile strength or elongation behavior upon applied stress. These types of analysis are however only applicable to macroplastic but were widely applied in the reviewed literature.

## 8. Leaching: Impact of UV exposure on release of plastic chemicals

When plastic particles come into contact with (sea)water or (marine) organisms, any organic and inorganic compounds, including additives and NIAS, have the potential to leach and as such become directly or indirectly bioavailable (Kühn et al., 2020; Hahladakis et al., 2018; Koelmans et al., 2014). The extent of such chemical leaching is determined by many intrinsic factors, including the polymer properties (e.g., type, crystallinity, surface area) and chemical properties (e.g., concentration, molecular weight, polarity, solubility), as well as extrinsic environmental parameters (e.g., temperature, pH, salinity) (Sørensen et al., 2024). For example, exposure to natural sunlight led to an increased release (by a factor of up to 5) of phthalates from PVC (cable) compared to dark conditions but had no impact on the phthalate release from PE bags (Paluselli et al., 2019). UV exposure may impact the aqueous leaching of chemicals, as well as influencing the potential toxicity of the leached chemical cocktail originating from marine plastic particles and debris. A range of approaches and techniques have been used to try and elucidate the chemical composition of plastic leachates during plastic aging. At the simplest level, dissolved and total organic carbon (DOC/TOC) has been measured as a proxy of leached chemicals in several studies (Chen et al., 2023; Wu et al., 2022; Zhu et al., 2020). However, the majority of studies have utilized targeted or non-targeted analysis of organics by GC-MS or liquid chromatography mass spectrometry (LC-MS). More recently, the use of optical characterization for the identification of leached plastic additives was explored and fluorescence signatures were described for diethylhexyl phthalate (DEHP) and bisphenol A (BPA) (Lee et al., 2020b).

UV exposure impacts plastic leachate composition through several mechanisms. First, UV-degradable chemicals, such as antioxidants and UV stabilizers (e.g., benzophenone) may undergo UV-degradation at the particle surface or immediately after release to surface water – resulting in their depletion from the leachate (Sørensen et al., 2021). However, this can also result in the formation of UV-degradation products, such as tris(2,4-di-tert-butylphenyl)phosphate from degradation of the antioxidant additive Irgafos 168 (Wu et al., 2020). Second, the leaching of UV-labile associated chemicals (e.g., phthalates, plasticizers, organophosphate esters and other flame retardants) can be facilitated and accelerated through UV degradation of the polymer, which increases the available surface area from which the chemicals can leach (Sait et al., 2021; Sørensen et al., 2021; Wang et al., 2020). Finally, UV degradation may lead to the formation of polymer degradation products that are
sufficiently small to leach into the aqueous matrix (e.g., terephthalic acid from PET, various long-chain carboxylic acids and other oxidation products of PE and PP) (Biale et al., 2022; Gewert et al., 2018; Sarno et al., 2021; Sørensen et al., 2021; Takahashi et al., 2022).

A small number of studies have analyzed and quantified the additive chemicals present in beached plastics, attempting to draw conclusions on the previous release of these additives into seawater or their degradation. For example, Rani et al. (2017a) found that the concentrations of most UV stabilizers and antioxidants were higher in new plastics compared to beached plastics, indicating their release or degradation during use or once exposed to the environment. Tanaka et al. (2020) analyzed one benzophenone (BP-12) and three benzotriazole (UV-326, UV-328, and UV-327) UV stabilizers in small and large fragments of beached plastics from Hawaii. While concentrations were generally lower in smaller fragments, suggesting an enhanced release of these compounds during fragmentation and weathering processes that increase the available surface area, the authors emphasized that notable amounts of the chemicals were still present even in these smaller fragments, which are more prone to be ingested by wildlife.

To date very few studies factor in plastic chemicals when addressing plastic environmental hazards (Delaeter et al., 2022). Currently published studies typically either quantify only a few selected target chemicals or simply quantify DOC/TOC increase during leaching. As full characterization of the plastic chemicals present in a material requires the combined use of several techniques (GC-MS, LC-MS and ICP-MS) and expert knowledge to tackle non-target screening data, it is hard to arrive at a harmonizable recommendation for quantification of the leaching of toxic plastic chemicals and degradation products during UV-exposure. However, it is suggested that this is a critical aspect in realizing more accurate and robust risk assessment of marine plastic pollution and requires attention moving forward.

## 9. Toxicity of UV-degraded microplastics and plastic chemical leachates

It is becoming increasingly acknowledged that plastic debris, especially nanoplastics and small microplastics, as well as their associated chemicals, might represent a hazard to the marine environment (Cole et al., 2011; Gunaalan et al., 2020; Pirsaheb et al., 2020). Recent research has focused on assessing the toxicological effects induced through exposure to plastic particles and their leachates, although elucidating between the two remains challenging. Available data concerning the toxic effects of nano- and microplastics are mostly derived from studies using direct exposure of organisms to pristine particles, often of a single polymer type, a very narrow size distribution, spherical morphology and poorly characterized from a chemical perspective. In contrast, toxicity studies using (partially) UV-degraded nanoplastic and microplastic test materials that more accurately reflect the types of particles found in the marine environment are very limited. The few available studies are reviewed here and summarized in Table 2, where all but one (Bejgarn et al., 2015) were conducted in 2020 or later.

The toxicity of plastic particles and/or leachates from plastics irradiated by UV was assessed using both in vivo or in vitro assays. A range of different aquatic model organisms were used across the studies, including prokaryotes and eukaryotes, as well as invertebrates and vertebrates. A range of acute and sublethal endpoints have been studied in vivo by exposure of partially UV-degraded plastic particles to invertebrates (e.g., sea urchin) and vertebrates (e.g., fish). For example, Wang et al. (2020) investigated the toxicological effects of partially UV-degraded versus non-UV-degraded PS microplastics on juvenile grouper (Epinephelus moara). The UV-degraded particles were generated by exposing the PS microplastics to UV irradiation in seawater for 60 days, which resulted in the formation of PS nanoplastics and the leaching of polymer fragments and additives, including high concentrations of dibutyl phthalate (DBP) and di-isobutyl phthalate (DIBP). Several differences between test groups were observed, including a higher hepatic

bioaccumulation of partially UV-degraded microplastics, possibly due to the reduced size of the UV-degraded particles (~39 µm versus ~56 µm). Furthermore, DBP and DIBP were detected in the liver of fish exposed to UV-degraded PS microplastics, but not in fish exposed to non-UV-degraded PS microplastics, suggesting that the UV exposure led to increased leaching of these additive chemicals and the corresponding increase in hepatotoxicity.

Although very little is known about the toxicity of (partially) UV-degraded (micro)plastics on marine organisms, a slightly larger number of studies have been conducted using freshwater species, including algae such as *Scenedesmus vacuolatus* (Rummel et al., 2022) and *Raphidocelis subcapitata* (Simon et al., 2021), invertebrates such as *Chironomus riparius* (Khosrovyan and Kahru, 2021), *Daphnia magna* (Khosrovyan and Kahru, 2022) and *Neocaridina palmata* (Klein et al., 2021a), and juvenile and adult vertebrates (fish), such as *Danio rerio* (Zhang et al., 2021) and *Oryzias latipes* (Coffin et al., 2018).

## 10. Toxicity of UV-degraded plastic chemical leachates

The toxicity of aqueous chemical leachates derived from plastic has been described for a variety of marine species (Bejgarn et al., 2015; Capolupo et al., 2021; Capolupo et al., 2020; Coffin et al., 2018; Delaeter et al., 2022; Gandara e Silva et al., 2016; Halsband et al., 2020; Li et al., 2016; Thaysen et al., 2018). Furthermore, the direct leaching of chemicals from plastic following plastic ingestion has also been demonstrated (Kühn et al., 2020). As discussed above, the complexity of the chemical leachate deriving from a specific plastic material may be expected to increase during UV-degrading conditions. Some recent studies have shown that UV-degraded leachates can be more toxic than non-degraded corresponding leachates (Rummel et al., 2022; Wang et al., 2020), although the overall picture may be more complicated. For example, Bejgarn et al. (2015) assessed and compared the toxicity of UV-degraded and non-UV-degraded plastic leachates generated from 8 different consumer products using the marine micro-crustacean *Nitocra spinipes*. There was no consistent trend in how leachate toxicity changed as a function of irradiation time, with four plastics becoming significantly more toxic after irradiation, two becoming significantly less toxic and two that did not change significantly. The study highlights the complexity in assessing the toxicity of plastic leachates and the dangers of drawing conclusions based on a limited number of test materials. The results suggest that the chemical profile of a 'pristine' plastic material can strongly influence the degree of leachate toxicity, as well as influencing whether subsequent exposure to UV irradiation changes the degree of leachate toxicity.

To the best of our knowledge, only two marine organism cell species (bacteria) have been used in in vitro studies to assess the impacts of UV-degraded aqueous leachates produced from plastic materials: *Aliivibrio fischeri* (Klein et al., 2021b) and *Prochlorococcus* (Sarker et al., 2020; Tetu et al., 2019). Prokaryotes such as the marine photosynthetic microorganism, *Prochlorococcus*, are highly abundant in the world's oceans and significantly contribute to global primary production (Partensky et al., 1999). Sarker et al. (2020) reported that the toxicity of leachates produced from HDPE plastic bags and PVC plastic matting was reduced by natural environmental weathering processes. Despite being weathered for up to 112 days, both materials continued to leach substances that negatively affected *Prochlorococcus* growth, photo-physiology and membrane integrity. Klein et al. (2021b) studied the toxicity of aqueous leachates derived from 12 different UV-degraded and non-UV-degraded plastic materials. In contrast to the findings of Sarker et al. (2020), Klein et al. (2021b) concluded that UV exposure of plastics significantly increased the toxicity of plastic leachates to *Aliivibrio fischeri* in in vitro tests.

While differences in the chemical composition of leachates before and after UV irradiation have been demonstrated, to date no studies have been able to pin-point the specific chemical components driving the observed toxicity before or after UV-exposure. Importantly, this
<table><caption>**Table 2**
Overview of studies using in vitro or in vivo tests to investigate the toxicological effects of UV-weathered plastics compared to pristine plastics.*</caption><thead><tr><th></th><th>Toxicological tests, model species and exposure time and concentration</th><th>Plastics</th><th>UV-exposure method and duration, leaching method</th><th>Analysis of microplastics and leachates</th></tr></thead><tbody><tr><td><strong>References</strong><br>Klein et al. (2021b)</td><td><strong>Model:</strong><br>Marine bacteria, <i>Alivibrio fischeri</i><br><strong>Assays:</strong><br>Microtox (baseline toxicity), ISO 11348-3, 2007<br><strong>Exposure concentration:</strong><br>Equivalent to 5.22–668 mg plastic.</td><td>PP, PET, PS, LDPE, PVC, Bio-PBS, SB pellets</td><td><strong>Weathering:</strong><br>UV irradiation (UV-C: 19.1 ± 2.3 W/m² and UV-A/B: 2.3 ± 0.3 W/m² (UV-A, including 0.25 W/m² of UV-B)) under laboratory conditions in dryness and a subsequent leaching period in ultrapure water (atmospheric weathering) or directly in water (aquatic weathering).<br><strong>Aquatic weathering:</strong> leaching for 24 h in 1.1 L of ultrapure water during UV-A/ B irradiation at 33.4 °C.</td><td>Leachates analyzed by LC-MS nontarget screening.</td></tr><tr><td>Sarker et al. (2020)</td><td><strong>Model:</strong><br>Two marine bacteria strains of <i>Prochlorococcus</i>: NATL2A and MIT9312<br><strong>Assays:</strong><ul><li>Membrane Integrity</li><li>Functional Chlorophyll a Fluorescence</li><li>Flow cytometry: concentration and relative chlorophyll fluorescence of cultures</li></ul><strong>Time points:</strong> 3, 24, and 48 h<br><strong>Exposure concentration:</strong></td><td>HDPE grocery bags and PVC matting</td><td><strong>Weathering:</strong><br>Natural tidal estuarine water for 17 and 112 days.<ul><li>17 days: 23.4–25.1 °C, 50000–50,400 mS/cm, 4.1–6.7 (UV index).</li><li>Additional period until 112 days: 13.6–21 °C, 49600–53,300 mS/cm, 0.9–5.1 (UV index)</li></ul><strong>Leaching:</strong><br>Weathered plastics were cut into ≈1 cm² pieces, placed into sterile artificial seawater, left for 5 days. Solutions were filtered through a 0.2 µm filter.</td><td>Analysis of leachates by ICP-MS and ICP-OES.</td></tr><tr><td>Tetu et al. (2019)</td><td><strong>Model:</strong><br>Two marine bacteria strains of <i>Prochlorococcus</i>: NATL2A and MIT9312<br><strong>Assays:</strong><ul><li>Flow cytometry (growth)</li><li>Functional chlorophyll a fluorescence measurement</li><li>RNA sequencing and analysis</li></ul><strong>Time points:</strong> 3, 24, 48 and 72 h<br><strong>Exposure concentration:</strong></td><td>HDPE grocery bags and PVC matting</td><td><ul><li>Weathered PVC dilutions: 1, 2, 10, and 50 % (equivalent to 0.5–25 mg/mL or 0.002–0.1 pieces/mL)</li></ul><strong>Weathering/leaching:</strong><br>Plastic pieces (~1–2 cm²) leached for five days under the following conditions: 22 °C in a shaking incubator at 100 rpm with continuous illumination of ~20–40 µmol photons m⁻² s⁻¹</td><td>Analysis of leachate by LC-MS, ICP-OES and ICP-MS</td></tr><tr><td>Bejgarn et al. (2015)</td><td><strong>Model:</strong><br><i>Nitocra spinipes</i> (micro-crustacean)<br><strong>Assays:</strong><br>Acute toxicity test<br><strong>Endpoints:</strong><br>LC10 and LC50 for plastics producing toxic leachates<br><strong>Time points:</strong> 0, 96, 192 and 288 h<br><strong>Exposure concentration:</strong> Liquid-to-solid (L:S) ratio of 10</td><td>Plastics:<br>PP, PS, LDPE, HDPE, PVC, PET, Bio-PET, PUR, PLA, mixed and sometimes unknown.</td><td><strong>Weathering:</strong><br>Plastics were ground, sieved and leached in natural brackish water (Baltic Sea, salinity = 7 ‰ ±0.2) for 72 h.<br><strong>Irradiation with artificial sunlight</strong><br>(300–800 nm, irradiance = 765 W/m²)</td><td>Chemical screening of leachates and detection: UPLC and qTOF-HRMS</td></tr><tr><td>Gewert et al. (2021)</td><td><strong>Model:</strong><br><i>Nitocra spinipes</i><br><strong>Assays:</strong><br>Acute toxicity test<br><strong>Endpoints:</strong><br>Average mortality (%)<br><strong>Time points:</strong> 96 h<br><strong>Exposure concentration:</strong></td><td>Pristine plastic pellets (PE, PP, PS and PET) and naturally weathered plastic collected from the marine environment: PVC, PET, PP, and two positive controls.</td><td><strong>Weathering:</strong><br>Plastic pellets suspended in floating mesocosms opened to the air in the Oslo fjord for 7 months.<br><strong>UV light exposure:</strong> 700 W/m² for 5 days (equivalent to an exposure of about 510 days to European mean solar irradiance)</td><td>No chemical analysis: assumption that there were no changes in production of the plastic products and that the initial chemical composition and the additives, consequently, were the same for both the old and new products.</td></tr><tr><td></td><td><ul><li>The highest concentration of the leachates: L:S ratio of 10.</li></ul></td><td></td><td></td><td></td></tr></tbody></table>

(continued on next page)
<table><caption>**Table 2 (continued)**</caption><thead><tr><th></th><th>Toxicological tests, model species and exposure time and concentration</th><th>Plastics</th><th>UV-exposure method and duration, leaching method</th><th>Analysis of microplastics and leachates</th></tr></thead><tbody><tr><td>Quade et al. (2022)</td><td><strong>Model:</strong><br>Sea urchin, <em>Paracentrotus lividus</em><br><strong>Assays:</strong><br>Embryo test<br><strong>Endpoints:</strong><br>Growth<br><strong>Time points:</strong> 0, 7, 28, 60 and 120 days<br><strong>Exposure concentration:</strong><br>3 dilutions (x1/10, x1/3, x1/1) of 1 g/L lixiviates of undegraded (day 0) and degraded (day 7, 28, 60 and 120) polymers (3 types) in the 3 habitats.</td><td>Commercial PE, PLA and PBAT plastic bags</td><td><strong>Weathering:</strong><br>2 × 15 cm pieces cut from each bag, tested in 3 mesocosm facilities under environmentally conditions: simulated pelagic (PEL), benthic (BEN) and littoral (LIT, sand system with no seawater) habitats.<br>PEL and BEN: flow-through-systems using natural 60 µm- filtered seawater. Conditions during 120 days: average temperature = 17.8 °C, photosynthetic active radiation: 0 to 800 µmol/m²s</td><td>In both biodegradable bags, additives containing an amide group were confirmed, but results were not published and detection methods not specified.</td></tr><tr><td>Tong et al. (2022)</td><td><strong>Model:</strong><br>Copepod, <em>Tigrionus japonicus</em><br><strong>Assays:</strong><br>Acute toxicity test<br><strong>Endpoints:</strong><br>Mortality<br><strong>Time points:</strong> 48 h exposure<br><strong>Exposure concentrations:</strong><br>• 20 and 500 mg/L of microplastics<br>• Mixture: 20 mg/L of microplastics and four concentrations of Triclosan (TCS): 90.50, 229.25, 349.16 and 732.08 µg/L</td><td>Biodegradable plastics from PBAT and PL, conventional plastics: PE</td><td><strong>Weathering:</strong><br>PE (thickness: ~0.03 mm) and PBAT/ PLA (thickness: ~0.04 mm) bag strips (50 × 20 mm) added to a 10 % hydrogen peroxide solution.<br>After 14 days, microplastics were collected, washed and dried at 60 °C, then passed through 400-mesh and 1400-mesh sieves to control the size from 10 to 40 µm.</td><td>None</td></tr><tr><td>Wang et al. (2020)</td><td><strong>Model:</strong><br><em>Epinephelus moara</em> (juvenile fish)<br><strong>Assays:</strong><br>Growth inhibition and hepatic pathological damage<br><strong>Endpoints:</strong><br>Growth, hepatic lipid homeostasis, hepatic cytotoxicity and gene expression responses responsible for the MP-induced hepatic lipotoxicity.<br><strong>Time points:</strong> 1, 6, 11, 16, 21, and 26 days<br><strong>Exposure concentrations:</strong><br>Fish exposed to seawater containing 1.5 and 15 mL/L of leachates, and fed with microplastics spiked diet: 2.0 and 20 mg/g dry feed</td><td>Commercial microbeads (C-PS), pristine fragments (P-PS), and UV photodegraded PS fragments (PD-PS)</td><td><strong>Weathering:</strong><br>• C-PS (mean diameter = 22.3 µm): in the ultrapure water stock solution<br>• P-PS (20–100 µm): prepared by mechanical grinding and subsequent sieving.<br>• PD-PS: from dried P-PS with filtered natural seawater and were UVA/B irradiated for 60 days (113 ± 45 W/m², 25 ± 3 °C).</td><td>Microplastic and nanosized particles morphology: SEM and TEM<br>Sample characterization: CLSM, ATRFTIR, X-ray photoelectron spectroscopy, EPR spectrometer and contact angle analyzer.<br>Plastic chemicals in leachates and livers: GC-MS</td></tr></tbody></table>

* Acronyms: polypropylene (PP), polyethylene terephthalate (PET), polystyrene (PS), low-density polyethylene (LDPE), polyvinylchloride (PVC), bio-based polyethylene succinate (Bio-PBS), starch blend (SB), polybutylene adipate terephthalate (PBAT), polylactic acid (PLA), polyurethane (PUR), liquid chromatography coupled to high-resolution mass spectrometry (LC-QTOF-MS), ultra-performance liquid chromatography (UPLC), Inductively Coupled Plasma-Mass Spectrometry (ICP-MS), Inductively Coupled Plasma-Optical Emission Spectrometry (ICP-OES), Micromass quadrupole time-of-flight high-resolution mass spectrometry (qTOF-HRMS), scanning electron microscopy (SEM), Transmission electron microscopy (TEM), Laser scanning microscopy (CLSM), Attenuated total reflection Fourier transformed infrared spectroscopy (ATRFTIR), Electron paramagnetic resonance (EPR), gas chromatography mass spectrometry (GC-MS).

study provides another good example of the importance of not attempting to assess plastic toxicity based on the polymer material alone.

## 11. Toxicity of biodegradable plastics

An interesting observation in a study by Zhang et al. (2021) is the earlier occurrence of toxicological effects for UV-degraded biodegradable plastics. It is proposed that the faster degradation in the environment of biodegradable plastics compared to conventional thermoplastics leads to an earlier release of toxic particulate and chemical degradation products. A recent study by Quade et al. (2022) demonstrated the correlation between the degree of UV-induced plastic

degradation and the toxicity of the corresponding leachates using commercial PE and biodegradable plastic bags. In the toxicity tests, sea urchin larvae (<em>P. lividus</em>) were exposed to cryomilled samples of different plastic bags allowed to degrade under natural environmental conditions for up to 120 days in outdoor mesocosms. Samples of each bag were placed in mesocosms simulating pelagic, littoral and benthic marine environments. No toxicity was observed for any of the PE bag samples. The initial toxicity of the biodegradable plastic bags was lost after 7 days incubation in the mesocosms, suggesting the toxicity of the pristine materials was primarily driven by readily leachable chemicals. The bags from the littoral zone of the mesocosms showed an increase in toxicity after 120 days incubation, which the authors suggest might be the result of UV-induced changes to the biodegradable plastics.
Together, these two studies suggest that (i) biodegradable plastics might contain different and more toxic additive chemicals than some thermoplastics, (ii) biodegradable plastics might be more toxic than conventional thermoplastics (at least PE) following UV degradation, and (iii) toxicity can be related to the degradation process and that the faster rate of degradation for biodegradable plastics may lead to higher concentrations of toxicants in the environment over shorter time scales. It is important to note, however, that 'bio'-associated plastics often comprise blends of several materials, making a toxicological evaluation difficult (Klein et al., 2021b). The conditions under which plastic additives may leach from plastics, and how this varies for different polymer types and different marine environmental conditions has only recently begun to be investigated.

## 12. Adsorption capacity of UV-degraded plastics

Studies have demonstrated the ability of plastic to adsorb organic and inorganic contaminants in the marine environment (Mato et al., 2001). These can include persistent organic pollutants (POPs), such as polyaromatic hydrocarbons (PAHs) or polychlorinated biphenyls (PCBs) (Hirai et al., 2011; Rios et al., 2007), as well as metals (Guo and Wang, 2019; Liu et al., 2022). The physicochemical properties of the plastic play an important role in pollutant adsorption, including size and shape (and therefore surface area), morphological and chemical surface properties, crystallinity, and zeta potential – all of which can be altered through degradation. In addition to the physicochemical properties of a plastic material, the adsorption of contaminants is influenced by environmental conditions (e.g., salinity, pH, temperature, DOM content) and the properties of the adsorbate (polarity, hydrophobicity, molecular weight) (Guo and Wang, 2019; Hartmann et al., 2017). Biofilms from the plastisphere have been identified to accumulate trace metals, though it remains currently unknown which influence the genetic or functional diversity of these biofilms might have on the adsorption of chemicals (Lenoble et al., 2024). The interactions between biofilm formation/biological attachment, adsorption of chemicals and UV degradation require further study, though it has been reported that biological attachment usually occurs on the side of a floating plastic fragment which is exposed to seawater, i.e. not facing sunlight (Hu et al., 2025).

This review identified four studies which examined the effect of UV aging in seawater on the adsorption capacity of plastics, two of which investigated adsorption of organic compounds and two that studied metal adsorption. Dhavamani et al. (2022) investigated the impact of salinity, temperature and UV irradiation (broad spectrum) over 5 days on the release and re-adsorption of phthalic acid esters (PAEs) with LDPE, HDPE, and a recycled PE material. While leaching rates and adsorption capacity were found to vary between polymers, exposure of the plastic materials to UV irradiation did not have a significant effect on adsorption, although it did increase the leaching of PAEs. Wu et al. (2020) studied changes in the adsorption capacity of 2,2',4,4'-tetramodiphenyl ether (BDE-47) to PS after a 90-day exposure to UV irradiation (254 nm) in air and filtered natural seawater. The effects of salinity, pH and dissolved organic matter (DOM) were also investigated. The authors observed a decrease in the adsorption capacity of partially UV-degraded PS relative to the pristine PS material. This effect was greater for polymers exposed to UV irradiation in air than those exposed in seawater, which was hypothesized to result from an increased appearance of functional groups containing oxygen on the particle surface.

Mao et al. (2020) studied the adsorption of zinc, copper, nickel, cadmium and lead to PS microplastics exposed to UV irradiation (340 nm) for ~2 months in air, artificial seawater and ultrapure water. FTIR and SEM analysis was used to determine the degree of UV-degradation between exposure conditions. Although the number of oxygen-containing functional groups were seen to increase with UV exposure time, there was no significant difference between UV exposure conditions (air, seawater, ultrapure water). In contrast, physical degradation

was observed to be the most advanced on PS exposed to UV irradiation in air. Increased metal adsorption was observed with increased exposure duration to UV irradiation. Han et al. (2021) investigated the effect of UV irradiation exposure (340 nm) and particle size on copper(II) adsorption to PE, PP, and PET in artificial seawater over 3 months. The authors observed variation in the adsorption capacity of copper(II) between the different polymer types, but all exhibited an increase in copper(II) adsorption after exposure to UV irradiation. Changes in surface chemistry were characterized by ATR-FTIR, which demonstrated the generation of oxygen-containing functional groups at the material surface of all three polymer types. Significantly, an additional peak was observed at 850 cm⁻¹ in the IR spectrum of PE exposed to UV irradiation in seawater, which the authors suggested was adsorption by -CCl₃ which had formed through substitution reactions at the polymer surface.

The above-mentioned studies investigated different combinations of adsorbates and polymer types, and used different degradation protocols, different seawater sources and different UV exposure wavelengths and times. This makes any direct and meaningful comparison between the four studies very challenging. For example, it may be argued that natural seawater more accurately reflects environmental conditions, but the influence of biodegradation can make it difficult to isolate the effects of UV degradation. Furthermore, Wu et al. (2020) noted that the formation of a biofilm on plastics aged in seawater may have contributed to a reduction in the adsorption of BDE-47. This effect was seen to a greater extent in dark controls, as the biofilm growth was limited under intense UV exposure conditions. The use of autoclaved or sterile-filtered seawater would eliminate biofilm formation and therefore exclude its influence on adsorption, allowing for observation of the impact of UV degradation alone.

The studies conducted by Mao et al. (2020) and Han et al. (2021) did not include dark controls, focusing only on trying to compare the partially UV-degraded polymer particles to their pristine analogues. Dark controls allow the impact of UV-degradation to be isolated from any potential effects of seawater exposure, especially in non-sterile conditions. Although UV irradiation has a substantial impact on plastic degradation, it is not the only viable degradation mechanism and studies have demonstrated degradation under dark conditions in various media, including seawater, highlighting the importance of including dark controls to elucidate the impact of UV (Da Costa et al., 2018). The adsorption studies reviewed herein highlight the need for standardized approaches to allow comparison between studies, as well as the necessity of dark controls to better isolate the impact of UV-degradation on contaminant adsorption. Finally, the studies reviewed here investigated organic and inorganic adsorption in isolation, however both classes of contaminants will be present as part of complex pollutant mixtures in most marine environmental compartments. Therefore, further studies may also consider the adsorption of organic and inorganic contaminants together, and how UV-degradation may impact this.

## 13. Identified knowledge gaps

Despite a recent increase in studies investigating the UV-degradation of plastic materials in the marine environment (Fig. 2), this review highlights that several key knowledge gaps remain. While there is increasingly strong evidence that UV-degradation of larger plastic items can lead to the formation of micro- and nanoplastics, the exact mechanisms of their formation remain ambiguous. The most recent studies indicate that significant fragmentation of large plastic items into smaller and smaller pieces is unlikely to be the primary result of UV-degradation. Instead, UV-degradation appears to result in physicochemical changes to the surface of the plastic material, which leads to the release of small micro- and nanoplastics through ablation-type mechanisms. It has been suggested that this is a continuous process, where the emitted particles leave behind a more 'pristine' surface on which the process starts over once more, gradually eroding the material away over long periods of time (Andrady et al., 2019). To better
understand the mechanism(s) of plastic UV-degradation and the subsequent formation and release of small particles, it is therefore recommended that future studies have a specific focus on following such surface changes and its impact on particle release to clarify this mechanism. Furthermore, the persistence and continued UV-degradation of the released micro- and nanoplastics, as well as the formation of molecular degradation products, needs more detailed study. While methods applicable to characterizing the size and surface changes of microplastic (>20 µm) throughout degradation processes exist and to a large extent are applied in the reviewed literature, there remains a lack of analytical techniques able to fully characterize and quantify particles <10 µm, particularly in the low concentrations expected from UV-studies in the lab or in situ. However, emerging techniques such as optical photothermal IR spectroscopy (O-PTIR) are suitable for certain matrices and could be considered in future plastic UV-degradation assessments.

Very few studies have looked at plastic degradation mechanisms as a simultaneous set of processes that influence each other. This is partly due to the longevity of plastic materials under natural weathering conditions: a study by Napper and Thompson (2019) found biodegradable and conventional plastic bags to remain functional (still able to hold >2 kg of groceries) after three years in soil and marine waters. Conducting laboratory-based UV-degradation studies that also investigate simultaneous mechanical and abrasive degradation mechanisms would provide a clearer understanding of the rate at which plastic degrades in the environment and the rate at which microplastics and nanoplastics are released. Similarly, combining UV-degradation studies with biodegradation studies would provide important knowledge about the persistence of particulate and chemical degradation products, as well as identifying whether UV-degradation helps to facilitate biodegradation through changes in the material surface chemistry. This highlights the need for controlled laboratory UV-degradation studies that mimic marine environmental conditions more closely, for example by increasing complexity through mesocosm setups, assessing UV-degradation under different marine environmental conditions, using consumer end-products as test materials and combining several degradation mechanisms concurrently. Such information is currently lacking, and difficult to produce robustly, but critical for fully understanding the environmental fate and behavior of plastic materials and their associated chemical and particulate degradation products.

Additive chemicals and NIAS have been shown to readily leach from many plastic materials and cause toxicity to marine organisms. While most available studies have focused on measuring and assessing plastic fragmentation and changes in the physicochemical properties of the material surface following UV-degradation, the leaching of additive chemicals and NIAS under UV exposure and the toxicity of chemical and particulate degradation products are rarely investigated. Additive chemical composition has the potential to be highly influential for controlling the rate and degree of UV-degradation, but the role of plastic additive chemicals in UV-degradation of plastic material in the environment has not been studied in sufficient detail. UV stabilizers are widely used in plastic consumer products, often being added to the polymer matrix during bulk production of plastic pellets. This class of additive chemicals are designed to specifically prevent or reduce the rate of plastic degradation and their presence in test materials should be known prior to use.

Furthermore, little data exists on the environmental persistence of the leached additive chemicals and NIAS, including their susceptibility to UV-degradation and biodegradation, and what degradation products might be formed from these leached chemicals. Such processes are important to understand from a hazard assessment perspective, where the toxicity of leached additives and NIAS may change due to UV-degradation. Toxicity has been shown to vary for different plastic types following UV-degradation, with some plastic products becoming more toxic, others less toxic, and some not changing significantly. The limited data available suggests that toxicity is highly dependent upon the specific additive chemical composition and concentration present in

the bulk polymer material. As with degraded plastic materials, different studies have reported more, less or no change in the toxicity of plastic leachates following UV exposure. Furthermore, the testing of plastic chemical leachates from one specific time point does not necessarily give a representative picture of the whole leaching process during the degradation of a plastic material in the marine environment. Further work is required to ascertain how UV-degradation of plastics and their leachates affects toxicity. Where viable, experimental set-ups should aim to include a comprehensive and interdisciplinary approach that allows toxicological assessment of plastics and leachates before, during and after the UV exposure period.

## 14. Recommendations for harmonized UV-exposure protocols for plastic and microplastic

Based on the findings from our review and critical assessment, Fig. 4 provides our best recommendation for a harmonized UV-degradation study methodology for plastic and microplastic that will allow comparability and read-across between different studies. Furthermore, Table 3 proposes a scheme for scoring the quality of selected analysis techniques for measuring the most important endpoints related to (micro)plastic UV-degradation, with the aim of providing reliable and relevant data for extrapolation to risk assessment. None of the studies reviewed herein would achieve maximum scores under the proposed scheme. This is primarily due to the fact that very few studies have focused on the role of additive chemical and NIAS leaching from plastic, and those that have often omitted a supporting analysis of material degradation and fragmentation. Of 63 studies evaluated in this review, only 10 investigated both these aspects.

## 15. Conclusion

Being able to robustly and accurately determine the risks associated with the partially degraded plastic materials that comprise most of the global marine plastic pollution is important for the development and implementation of informed and impactful mitigation actions. The knowledge gaps outlined by the current study are important for understanding the risks associated with plastic materials undergoing UV-degradation processes in the marine environment.

Based on a careful review of existing practices and insight into availability of analytical techniques to characterize and quantify plastic and microplastic particles, we arrive at a recommendation for harmonized practice in exposure design and an optimal selection of combined assessment methods. For data to be applicable in marine environmental risk assessments, we emphasize the need for an accurate quantification of fragmentation and release of micro- and nanoparticles as well as a broad analysis of leaching additives and degradation products that may be harmful to marine life. Although this review has focused on studies of degradation in the marine environment, we believe these recommendations are directly transferable to studies in other matrices (e.g., freshwater in particular).

There is very limited knowledge and data regarding the toxicity of partially degraded plastic materials and degradation products towards marine organisms. While little is currently known about the potential toxicity of UV-weathered plastics and plastic leachates relative to their pristine forms, the limited data available suggests a complex picture. This makes it very challenging to be able to draw any strong conclusions or assess differences in risk between degraded and non-degraded plastic materials comprised of the same polymer type.

Understanding which polymer types and additive chemicals lead to greater or lesser toxicological responses following UV irradiation should be a priority in order to develop mitigation actions that can reduce the environmental impact of plastic pollution, for example by implementing a replacement approach for the polymers and additive chemicals identified as the most toxic and persistent in the marine environment (i.e., those representing the highest risk). Making the complete datasets
<figure><img src="image_6.png" /><figcaption>Fig. 4. Recommendations for future harmonized experimental protocols to study the mechanisms and effects of UV-degradation of plastics under aqueous conditions.</figcaption></figure>
<table><caption>**Table 3**
Proposed scoring system for UV-degradation characterization and quantification methods.</caption><thead><tr><th></th><th>Surface changes</th><th>Fragmentation</th><th>Leaching</th><th>Bulk material properties</th></tr></thead><tbody><tr><td>Low score</td><td>Visual</td><td>Weight loss</td><td>Total or dissolved carbon</td><td>Recommended to use several for full score.</td></tr><tr><td>Medium score</td><td>FTIR/Raman/ hyperspectral or SEM</td><td>Optical microscopy or SEM images</td><td>Target analysis of one group of compounds</td><td>Thermal properties (DSC, TGA), crystallinity, tensile strength, microhardness, molecular size</td></tr><tr><td>High score</td><td>FTIR/Raman/ hyperspectral and optical microscopy/ SEM</td><td>Size distribution analysis of particles</td><td>Target analysis of multiple compounds or screening with one technique</td><td></td></tr></tbody></table>

available will be crucial to accelerate these processes.

### CRediT authorship contribution statement

**Laura M. Hernandez:** Writing – original draft, Visualization, Validation, Methodology, Investigation, Formal analysis, Data curation, Conceptualization. **Lucy Howarth-Forster:** Writing – original draft, Visualization, Validation, Methodology, Investigation, Formal analysis, Data curation. **Lisbet Sørensen:** Writing – original draft, Validation, Supervision, Methodology, Investigation, Funding acquisition, Conceptualization. **Andy M. Booth:** Writing – original draft, Validation, Methodology, Investigation, Funding acquisition, Conceptualization. **Alice Vidal:** Writing – original draft, Visualization, Validation, Methodology, Investigation, Formal analysis, Data curation. **Nathalie Tufenkji:** Writing – original draft, Validation, Supervision, Project administration, Methodology, Investigation, Funding acquisition. **Richard Sempéré:** Writing – original draft, Validation, Supervision, Project administration, Funding acquisition, Conceptualization. **Natacha Schmidt:** Writing – original draft, Visualization, Validation, Project administration, Methodology, Investigation, Formal analysis, Data curation, Conceptualization.

### Declaration of competing 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

Data will be made available on request.

### Acknowledgements

This research was conducted as part of the JPI Oceans ANDROMEDA project, which received funding through the Research Council of Norway (Grant numbers 312262 and 311313/E40) and the French Agence Nationale de Recherche (ANR-19-JOCE-0002-01). AB and LS were also supported through the Research Council of Norway project REVEAL (Grant number 301157). LMH and NT acknowledge the Canada Research Chairs program, the Natural Sciences and Engineering Research Council of Canada, and the Killam Research Fellowship. LHF was supported through the Natural Environment Research Council (NERC) via the United Kingdom Research and Innovation (UKRI) to the Centre for Doctoral Training in Sustainable Management of UK Marine Resources (CDT SuMMeR) (Grant number NE/W007215/1).

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