and to meet regulations for food packaging. Some concentrations can be lower than the detection limits; and nondetects should never be ignored and were included as zeros while calculating the average additive concentrations, as shown in Figure 2.4.
A second reason why additives are in lower concentrations than expected is prior leaching from plastic into the environment or degradation of the additive (Rani et al. 2017b; Tanaka et al. 2020). The third reason is some of the additives detected in marine plastic debris could be adsorbed from the surrounding environment rather than being intentionally added. Plastic polymers are routinely used as passive samplers to monitor environmental pollutants in water, because they are excellent at adsorbing compounds from the surrounding environment (Koelmans et al. 2016). Because of this difficulty in determining the source of additives, the use of plastic debris samples to assess global spatial and temporal trends is complicated. Even so, Prunier et al. (2019) noted that mesoplastics from English coastal areas (Massos and Turner 2017; Turner and Solman 2016) or Chinese littoral areas (Wang et al. 2017) had element concentrations in the same order of magnitude as those from the open ocean (Prunier et al. 2019). This differs for organic pollutants, whose concentrations are greater in plastic debris from coastal areas than from the open ocean (Hirai et al. 2011). These findings are likely driven by the global distribution of naturally occurring elements in seawater and point sources influx from human activities in coastal regions (Net et al. 2015).
2.5.2 Abiotic Samples
Water and sediment are more frequently sampled than biota in plastic additive studies. For organic additives the sediment concentrations are routinely two to three orders of magnitude greater compared to that in water (Figure 2.5), as expected based on their log K ow or sediment water partition coefficients.
Monitoring surface water for additives is particularly well suited for assessing global spatial and temporal comparisons, because many plastic additives are soluble enough in water to be detectable by standard monitoring techniques. Also, seawater is a globally accessible sampling matrix that is regionally influenced by local point sources. By performing a meta‐analysis of phthalate concentrations in surface waters, Berge et al. (2013) reported that the European and Chinese coastal waters had higher median DEHP concentrations of approximately 1 μg/L, compared to North American waters (approximately 0.3 μg/L). However, an updated analysis of DEHP in marine waters only shows that the range of means in Asia are relatively higher than America, followed by Europe (Figure 2.6). The variability within each continental region, and the differences in sampling times, are too large to make definitive spatial comparisons. Berge et al. (2013) observed temporal trends in the phthalate concentrations in European fresh and seawaters, which were increasing until around 2000, then decreasing due to regulatory influence (Berge et al. 2013). The marine surface waters do not reflect this trend, but no studies before 2000 were included (Figure 2.6). Zhang et al. (2018b) found a vertical distribution of phthalates in seawater with greater concentrations at the surface of the ocean that decreased slightly with depth until increasing near the bottom.
Figure 2.5 Range of concentrations of three plastic additive classes measured in paired seawater and sediment samples. Note: Data were taken from de los Rios et al. (2012) for 4‐nonylphenol and bisphenol A in natural samples from the Bay of Biscay; and from Zhang et al. (2018b) for phthalates in Bohai and Yellow Seas.
Sediment cores can offer an in situ sample archive for environmental monitoring programs to signal when regulations were needed and also to track changes in pollutant levels resulting from regulations or societal shifts. Radionuclide dating is especially valuable in determining when different sediment layers with additives were deposited. Five studies have used sediment coring methods to understand temporal trends in plastic‐related chemical concentrations in Asian coastal waters only (Hashimoto et al. 2005; Moon et al. 2009; Peng et al. 2007; Wang et al. 2018; Zhang et al. 2013). Three of these studies focused on APs and/or BPA, revealing temporal trends that could be explained by local/regional land‐use or policy changes. Wang et al. (2018) observed increasing concentrations of plastic additive phenolic antioxidants, such as BHT, from the bottom to the top of cores collected close to the coast, indicating increasing temporal trends of these compounds in Chinese marine habitats. Zhang et al. (2013) explained complex periodic spikes in HBCDs measured in layers of a sediment core by China’s national economic development initiatives and the opening of a nearby corporation in 2005.
2.5.3 Biotic Samples
Within biota studies, plastic additives were detectable in tissues from at least 134 species. The first sampling began before 1978 (Giam et al. 1978), and most studies have focused on fish, bivalves, and other invertebrates with a few studies addressing plants, turtles, birds, and mammals (Figure 2.7).
Attempting spatial and temporal comparisons in similar species for a particular compound class, other than PBDEs and HBCDs, is difficult because the published data is sparse. Past reviews have shown elevated levels of HBCDs in samples near chemical production or application facilities, indicating that some plastic additives are released directly into the environment instead of being leached from plastic products (Covaci et al. 2006). Several temporal trends show increasing BFR concentrations until approximately 2000, reflecting usage of the compounds in developed countries (Law et al. 2014). The continual increasing trends observed in the Arctic indicate their transportation to polar regions. For the other additive classes, fish offers the best sample type, but data are very limited after filtering the data for similar trophic level, habitat type, tissue analyzed, particular chemical reported in the same or convertible units, and summary statistic (e.g., mean or median). Filtering criteria are critical because additive concentrations can vary widely among fish species even from the same location (Gu et al. 2016), and additives do not distribute evenly throughout the body (Barboza et al. 2020). Filtered data available for comparison often include <20 individual fish from two or three locations on the global map and at different snapshots in time (Figure 2.8). It is not advisable to make global spatial generalizations with data like these.
Figure 2.6 An updated spatial comparison of mean concentrations of DEHP measured in marine surface waters from Asia, America, and Europe.
Figure 2.7 Distribution of plastic additive studies assessing different taxonomic groups.
Figure 2.8 Spatial comparison of bisphenol A mean concentrations in muscle tissue from
