Release kinetics and bioaccessibility of heavy metal from antifouling paint
particles in simulated digestive fluids: An emerging threat to marine biota
1.Introduction：
Biofouling remains a persistent and serious issue in the marine environment, posing significant challenges to various sectors, including shipping, aquaculture, and marine infrastructure. Marine organisms can colonize on the surfaces of submerged objects, leading to increased hull resistance, pipeline blockages, accelerated metal corrosion, and damage
to aquaculture facilities (Schultz et al., 2011; Liu et al., 2022). These impacts not only compromise the efficiency and safety of maritime op- erations but also result in substantial economic losses. According to incomplete estimates, the global economic loss attributed to marine biofouling reaches $1.5 billion annually (Jin et al., 2022).
Antifouling paints have been considered one of the most effective methods for preventing marine biofouling (Lindholdt et al., 2015).
- Correspondence to: H-X. Li, South China Sea Institute of Oceanology, Chinese Academy of Sciences, Guangzhou 510301, China.
** Corresponding author.
E-mail addresses: hxli@scsio.ac.cn (H.-X. Li), xuxr@gxu.edu.cn (X.-R. Xu).
https://doi.org/10.1016/j.marpolbul.2025.118596
Received 22 May 2025; Received in revised form 15 July 2025; Accepted 12 August 2025 Available online 18 August 2025
0025-326X/© 2025 Elsevier Ltd. All rights are reserved, including those for text and data mining, AI training, and similar technologies.
However, these paints can be damaged during their services, leading to the release of smaller particles known as antifouling paint particles (APPs) (Gaylarde et al., 2021). Data from the Organization for Economic Co-operation and Development indicated that approximately 6 % of a ship's coating mass is lost over its lifetime, with losses occurring during the painting process (1.8 %), through weathering during use (1 %), and during maintenance (3.2 %) (OECD, 2009). APPs have been widely detected in seawater, sediments, the atmosphere, and even within ma- rine organisms (Turner, 2021). The increasing release of APPs is now recognized as a significant and emerging source of contamination in marine environments (Compa et al., 2022; Sparks and Awe, 2022).
Antifouling paints often incorporate toxic chemical antifouling agents, such as heavy metals (e.g., copper, zinc) and organic compounds (e.g., copper oxide, organotin). (Amara et al., 2018). These substances can leach into marine environment, posing toxic risks to non-target or- ganisms (Li et al., 2023; Weber and Esmaeili, 2023). For instance, tributyltin (TBT) was widely used in antifouling paints but was found to be extremely harmful to marine organisms. Even at very low concen- trations (e.g., 20 ng/L), TBT can cause developmental toxicity in marine organisms and threaten the reproductive capabilities of fish and shellfish (Buskens et al., 2013). This has led to a global ban on TBT and high- lighted the urgent need for the development of environmentally friendly antifouling materials (Beyer et al., 2022). Heavy metal-based antifouling agents, such as copper and oxide, pyrithione copper, pyrithione zinc, have emerged as alternatives to organotin compounds (Kyei et al., 2020). However, the ecological risks associated with the release of heavy metals from antifouling paints into the marine environment remain inadequately assessed (Ali et al., 2020; Zhang et al., 2023a).
The heavy metals and toxic chemicals in APPs are not easily degraded and can accumulate in the marine environment (Turner, 2021). High concentrations of heavy metals are frequently detected in the seawater, sediments, and organisms near coastal docks, shipyards, and harbors, which may be attributed to the application and release of marine antifouling paints (Turner et al., 2017). APPs have been found rich in heavy metals and considered as a significant source and carrier for the transport of heavy metals in the marine environment (Muller- Karanassos et al., 2019; Gaylarde et al., 2021). Furthermore, APPs have been detected in the digestive tracts of a wide range of marine organ- isms, such as polychaete, fish, and seabird (Forero-Lopez et al., 2024). Notably, the direct ingestion and digestion of APPs rich in heavy metals by marine organisms is a crucial pathway for the bioaccumulation of heavy metals. These heavy metals can be transferred and amplified through the food chain, eventually affecting higher trophic level or- ganisms, including humans (Finkelstein et al., 2003; Muller-Karanassos et al., 2019). Additionally, studies have revealed that the massive release of copper ions can lead to the death of algae, affect the embry- onic development of carp, salmon, shrimp, crabs, and various mollusks, and destroy their blood cells (Forero-Lopez et al., 2024). However, there is still a significant gap in understanding the release behavior of heavy metals from APPs and their bioaccessibility in marine biota.
With the rapid development of marine economic industries, the production and usage of antifouling paints have also increased. The pollution and toxicity risks associated with APPs and antifouling agents in the marine environment are matters of concern for the safety of the marine ecosystem and human health (de Campos et al., 2022). To address these concerns, it is essential to investigate the release behaviors and bioaccessibility of heavy metals from APPs. Biological digestive fluids serve as an effective method for exploring the release of con- taminants within organisms and evaluating bioaccessibility (Holmes et al., 2020; Turner et al., 2008). Therefore, this study employed simu- lated digestion fluids to explore: (1) the release kinetics of heavy metal from APPs in different types of digestive fluids, such as those of poly- chaete, fish, and seabird; (2) the key factors influencing the release behaviors of heavy metals from APPs; and (3) the bioaccessibility of heavy metal released from APPs. Understanding these processes is crucial for developing strategies to mitigate the environmental impacts
of APPs and for the development of safer antifouling technologies.
- Material and methods：
2.1. Preparation of APPs in different sizes
A typical commercial antifouling paint with Cu2O as the primary biocide was purchased from Qi Jia paint company (Zhejiang, China) based on its large market share. This paint is an acrylic resin-based, tin- free and self-polishing paint. The preparation of APPs was conducted in accordance with the methodology described in a previous study (Simon et al., 2021). Briefly, the antifouling paint was brushed onto a clean blotting paper (without any metal additives) and allowed air drying in a fume hood. Afterward, the antifouling paint films were gently removed from the blotting paper and then cut into small pieces using sterilized ceramic knife and ground into particles under liquid nitrogen using a porcelain mortar and pestle. These paint particles were further pro- cessed by sieving through nylon mesh screens to isolate particles within different size ranges of 0.02–0.1, 0.1–0.5, 0.5–1.0, 1.0–1.5, and 1.5–2.0 mm. The predominant size range of paint particles in the environment is between 0.1 and 0.5 mm (Supplementary materials, Table S1).
2.2. Preparation of UV-aged APPs
The artificial aging of APPs used a UV irradiation chamber to emulate the photodegradation effects of sunlight, with an aim of obtaining the environmental aging characteristics of APPs. Exposure to UV irradiation can cause the cleavage of chemical bonds within APPs. This, in turn, alters their physical and chemical properties and poten- tially accelerating the release of additives, such as heavy metals. The irradiation chamber was equipped with eight UVB lamps, each emitting a radiation intensity of 20 mW.cm − 2 at a wavelength of 313 nm. APPs within the size of 0.1–0.5 mm were selected and evenly distributed on clean glass petri dishes, exposing to irradiation for a period of 7, 15, and 30 days to simulate varying degrees of aging (Simon et al., 2021). Throughout the irradiation process, APPs samples were manually turned daily within the chamber to ensure uniform exposure.
Infrared spectra of both pristine and aged APPs were acquired using attenuated total reflectance Fourier-transform infrared (ATR-FTIR)
spectroscopy with a Thermo Scientific Nicolet iS50R instrument (USA). Five replicates of APPs from different UV irradiation durations (0, 7, 15, and 30 days) were randomly selected for analysis. The ultramicroscopic morphology and elemental composition of both pristine and aged APPs were examined using scanning electron microscopy (SEM) with a Zeiss Gemini instrument (Germany), complemented by energy-dispersive X- ray spectroscopy (EDS) using an Oxford Inca Energy system (UK). Comparative analysis of the surface characters between pristine and aged APPs was conducted (Supplementary materials, Figs. S1-S3).
2.3. Preparation of simulated digestive fluids
To assess the release of heavy metals from APPs within marine or- ganisms, we selected three types of digestive fluids representative of polychaete, fish, and seabird. The design and preparations of these simulated digestive fluids (Supplementary materials, Table S2) fol- lowed by the established protocols (Turner, 2018; Coffin et al., 2019b). Artificial saltwater was prepared using ultrapure water and commercial salt (Instant Ocean® Spectrum Brands, USA) to achieve a salinity of 32 PSU at a pH of 7.0. The reported body temperatures of polychaete, fish, and seabird are 18 ◦ C, 25 ◦ C, and 40 ◦ C, respectively (Coffin et al., 2019a; Coffin et al., 2019b). Therefore, the experimental temperatures were set to match these physiological conditions. To simulate poly- chaete digestion, sodium taurocholate (15.5 mM) and bovine serum albumin (BSA) (5 g.L − 1) were incorporated into the artificial saltwater, and the samples were incubated at 18 ◦ C. For the simulated fish digestive fluid, pepsin A (Sigma-Aldrich, Germany) was dissolved in artificial
Z.-L. Li et al. Marine Pollution Bulletin 221 (2025) 118596
seawater at a concentration of 2 g.L − 1, and the pH was adjusted to 2.0 using 1 M HCl and NaOH. The simulated seabird digestive fluid was prepared by dissolving 10 g of pepsin (Sigma-Aldrich, Germany) in 1 L of 0.1 M NaCl solution, with the pH adjusted to 2.5 using 1 M HCl and NaOH.
2.4. Heavy metal release experiments
A 0.1 g of pristine APPs within 0.1–0.5 mm was introduced into a 250 mL glass bottle, followed by the addition of 100 mL of simulated digestive fluid, resulting in a final solid-to-liquid ratio of 1 g.L − 1. Control treatments consisted of either seawater or a 0.1 M NaCl solution. The release of heavy metals from APPs was evaluated under the following various conditions: polychaete digestive fluid versus seawater control, fish digestive fluid versus seawater control, and seabird digestive fluid versus 0.1 M NaCl solution control. The different controls were used to simulate the specific ionic strengths and chemical compositions of the natural environments these organisms inhabit (Turner, 2018; Coffin et al., 2019b). Compared to the residence times of natural foods within organisms, the egestion of polymers seems to be processed more slowly. It is estimated that microplastics are retained in the digestive tracts of seabirds (Fulmarus glacialis) for several days to weeks (van Franeker et al., 2011). Consequently, APPs samples were placed into a shaking incubator and agitated at 130 rpm for a total of 144 h to better elucidate the kinetics of heavy metal release in simulated digestive fluids over an extended period. Leachate of the APPs was collected and filtered through a 0.45 μm cellulose acetate membrane at time points of 0, 0.5, 1, 2, 4, 8, 12, 24, 48, 72, 96, 120, and 144 h. The filtered samples were stored at 4 ◦ C for subsequent analysis, and each experiment was con- ducted in triplicate to ensure reproducibility.
2.5. Influencing factors on the release of heavy metals
Considering APPs within various sizes occurred in marine environ- ment and their particle sizes may influence the release of additive, this study examined the release of heavy metals from pristine APPs at different size ranges (0.02–0.1, 0.1–0.5, 0.5–1.0, 1.0–1.5, and 1.5–2.0 mm). Furthermore, APPs are particularly vulnerable to ultraviolet (UV) irradiation, which can induce weathering on their surfaces. This study utilized aged APPs (UV irradiation for 7, 15, and 30 days) to assess their effects on the leaching behaviors of heavy metals. The experimental design involved preparing 0.1 g of APPs within 100 mL of different simulated digestive fluids and subjecting them to a 24-h shaking period to mimic the average gastrointestinal transit time of food in marine organisms (Clements, 1997; Grigorakis et al., 2017; Holmes et al., 2020). Each experiment was performed in triplicate, with artificial seawater or a 0.1 M NaCl solution serving as the controls to compare against the simulated digestive fluids.
2.6. Analysis of heavy metals
A total of 0.01 g APPs were weighed and transferred into a Teflon digestion vessel. Acid digestion was performed using a mixture of hy- drochloric acid (HCl), nitric acid (HNO3), hydrofluoric acid (HF), and hydrogen peroxide (H2O2) in a volumetric ratio of 3:1:1:1. The digestion was carried out in a microwave oven (Anton Paar Multiwave PRO 41HVT56, Austria). Post-digestion, the vessels were transferred to a heating block set at 180 ◦ C to evaporate the acid mixture nearly to dryness, leaving approximately 1 mL of digestate. Both the APP diges- tion solutions and the leachates from the simulated digestive fluids were subsequently diluted with 2 % nitric acid. Heavy metal contents and their concentrations in the solution were determined using an induc- tively coupled plasma mass spectrometer (ICP-MS).
2.7. Quality assurance (QA) and quality control (QC)
All chemicals used in this study were guarantee reagent (GR) and purchased from ANPEL Laboratory Technologies Co., Ltd. (Shanghai, China). Glassware was pre-cleaned by soaking in a 20 % (v/v) nitric acid solution for 24 h and subsequently rinsed with the ultrapure water multiple times. All samples were conducted in triplicate, and procedural blank controls were incorporated throughout the analytical process to ensure analytical accuracy and precision. The final concentrations of heavy metals were corrected by subtracting the blank values. A quality control sample composed by muti-element standard solution (GBW (E) 80,040) was analyzed every 10 samples, with measured heavy metal recoveries of 90–110 % per batch. Relative standard deviations (RSDs) for the replicate samples were less than 5 % in all cases. Method detection limits (MDLs) for Cu, Zn, Cd, Pb, Fe, Al, Mn, Cr, and Ni in APPs were 0.008, 0.160, 0.015, 0.070, 3.00, 6.50, 0.011, 0.070, and 0.003 μg.g − 1, respectively. In contrast, the MDLs for Cu, Pb, and Zn in three digestive fluids, seawater, and NaCl solution were 0.12, 0.07, and 0.01 μg.L − 1, respectively.
2.8. Release kinetic analysis of heavy metal from APPs
Three kinetic models, including the first-order kinetic model, second- order kinetic model, and diffusion-controlled model, were used to elucidate the release behaviors of heavy metals in various experimental conditions according to the literatures (Masset et al., 2021; Holmes et al., 2020; Coffin et al., 2019a).
The first-order kinetic model is expressed as follows:
ln ( Ceq − C ) = lnCeq − k1t (1)
The second-order kinetic model is expressed as follows:
The diffusion-controlled model is expressed as follows:
兀
The diffusion-controlled model based on Fick's second law for the dissolution of a semi-infinite slab can be described as follows (Ruby et al., 1992).
The quasi-equilibrium concentration of heavy metal in solution is denoted as Ceq, and C represents the concentration at a specific time point during release. The first-order kinetic rate constant is designated as k1, and the second-order kinetic rate constant is designated as k2. In the diffusion-controlled model, k3represents the parabolic diffusion rate constant, which is determined from the linear regression analysis of C versus t1/2. The term b is defined as the effective initial concentration. These kinetic models were applied to elucidate the release behaviors and mechanisms of heavy metals from APPs in digestive fluids (Supple- mentary materials, Tables S3-S5).
2.9. Bioaccessibility of heavy metals released from APPs
The bioaccessibility of released heavy metals was determined by calculating the ratio of heavy metal concentrations extracted by the digestive fluid to the concentration of heavy metals in the APPs (Holmes et al., 2020). Specifically, bioaccessibility (BA) can be calculated using the following formula:
Where C represents the concentration (mg.L − 1) of heavy metals in the digestive fluid at the end of a 24-h shaking period; V represents the
Z.-L. Li et al. Marine Pollution Bulletin 221 (2025) 118596
volume of digestive fluid (100 mL); MAPPs represent the mass of APPs (0.1 g); and CAPPs is the concentration (mg.g− 1) of heavy metals in APPs. This method provides a quantitative measure of the fraction of heavy metals that can potentially be released from APPs and become available for absorption in the gastrointestinal tract, which is crucial for risk assessment.
2.10. Statistical analysis
The normality of the data was first evaluated using the Shapiro-Wilk test, while the homogeneity of variances was assessed using the Levene test. Subsequently, if the data was normally distributed and exhibited homogeneity of variance, a one-way analysis of variance (ANOVA) coupled with Tukey's post hoc Honest Significant Difference (HSD) test was employed to determine the significance of the differences among the groups. In cases where the data did not meet the assumptions of normality or homogeneity of variance, the Mann-Whitney Utest and the Kruskal-Wallis H test were selected for significance testing. The statis- tical significance was set atp < 0.05. Data analysis was performed using Office 2021, IBM SPSS Statistics version 22, and Origin 2022 software.
