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JMSEJournal of Marine Science and Engineering
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  • Review
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16 November 2025

Organophosphate Esters in Marine Environments: Source, Transport and Distribution

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1
National Research Center for Geoanalysis, Chinese Academy of Geological Science, Beijing 100037, China
2
Beijing International Center for Gas Hydrate, Peking University, Beijing 100871, China
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Authors to whom correspondence should be addressed.
J. Mar. Sci. Eng.2025, 13(11), 2162;https://doi.org/10.3390/jmse13112162
This article belongs to the Section Marine Environmental Science
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Abstract

Organophosphorus esters (OPEs), widely utilized as flame retardants and plasticizers, are physically incorporated into those products and exhibit semi-volatility, resulting in release throughout their lifecycle. The ocean serves as a significant sink and plays a pivotal role in the global distribution and environmental fate of OPEs. However, the OPEs’ behavior and ecological effects in marine systems are not well understood. This review systematically examines recent advances in the sources, transport pathways, transformation mechanisms, and distributions of OPEs in the marine environment, and it also addresses current research limitations and suggests directions for future work. It is found that OPEs predominantly enter the marine environment through terrestrial input and in situ release; the transportation means include river input, long-range atmospheric transport, air–sea exchange, and oceanic circulation; and the degradation processes of OPEs are recognized as hydrolysis, photodegradation, and biodegradation. The distributions of OPEs in marine environments vary in different media, with their concentrations observed to range from pg m−3 to ng m−3 in marine air, ng L−1 to hundreds of ng L−1 in seawater, and pg g−1 dw to ng g−1 dw in sediments. The distributions of different species of OPEs are affected by many factors, such as compound properties, environmental conditions, and policy regulations. Comparisons between different regions and different seasons need to be further studied, and predictive models should be developed to better assess ecological risks and exposure pathways of OPEs.

1. Introduction

Organophosphate esters (OPEs) are a group of synthetic chemicals characterized by the general structure of O=P(OR)3 [1]. These compounds have been widely utilized as flame retardants and plasticizers [2]. OPEs were previously regarded as an ideal substitute for brominated flame retardants (BFRs) [3] due to their simpler production processes, lower cost, effective flame retardant performance, and good durability [4]. In recent years, with global restrictions on BFRs [5,6], OPE usage has surged from approximately 0.68 million tons in 2015 to over 100 million tons in 2018 [7].
Based on substituent differences, OPEs are typically categorized into alkyl-, aryl-, and chlorinated-OPEs [8]. Common alkyl-OPEs include triethyl phosphate (TEP) and tri-n-butyl phosphate (TnBP); typical aryl-OPEs include triphenyl phosphate (TPhP) and tricresyl phosphate (TCP); major chlorinated OPEs include tri(2-chloroethyl) phosphate (TCEP), tri(2-chloroisopropyl) phosphate (TCPP), and tri(1,3-dichloroisopropyl) phosphate (TDCPP) [3,4,9]. These compounds generally exhibit high chemical stability, moderate water solubility, and a relatively wide range of octanol–water partition coefficients (lgKow) [10,11], as detailed in Table 1. Their environmental migration and distribution behaviors vary due to differences in physicochemical properties [8,12]. For example, the water solubility of OPEs generally decreases with increasing molecular weight [11]. And chlorinated OPEs tend to exhibit stronger hydrophobicity and adsorption properties, making them more prone to accumulation in sediments and organisms. Trimethyl phosphate (TMP) and TEP possess high water solubility, whereas TCIP is virtually insoluble in water [13]. Most OPEs exhibit lgKOW values greater than zero, reflecting their hydrophobic nature and tendency to adsorb onto soil and sediments or accumulate within organisms [14,15,16,17]. Significant variations in vapor pressure exist among different OPEs. For instance, TEP and TCPP exhibit high vapor pressures, enhancing their volatility and potential for atmospheric entry, where they may subsequently adsorb onto particulate matter [8,18], enabling long-range transport. Table 1 summarizes the structures, names, abbreviations, and physicochemical properties of commonly studied OPE compounds.
In applications, chlorinated OPEs are primarily used as flame retardants due to their excellent fire-retardant properties [18], finding widespread use in building materials, insulation foams, and electronic components. Non-chlorinated OPEs are typically employed as plasticizers to enhance the flexibility of plastics and polymers. Additionally, OPEs serve as defoaming stabilizers, floor polishes, lubricants, varnish additives, and hydraulic fluid additives [8,19]. Notably, most OPEs are incorporated into products through physical mixing rather than chemical bonding [9,20] and exhibit semi-volatile properties. Consequently, OPEs readily migrate into environmental media via volatilization, leaching, and abrasion [14,21,22], leading to increasing detection rates in the environment. To date, OPEs have been detected in diverse environmental media including air, water, sediments, soil, and biological organisms [2,6,14]. These findings have raised concerns about the potential toxicological impacts of OPEs on ecosystems and human health [23].
The marine environment serves as a significant sink for OPEs. Current research indicates that OPEs can enter the ocean through multiple pathways, including river input [24,25,26], atmospheric deposition [27,28,29], and in-suit plastic degradation [30]. For instance, chlorinated OPEs, due to their strong hydrophobicity, are more likely to adsorb onto particulate matter and eventually enter sediments and organisms [3]; while certain highly volatile alkyl OPEs can affect distant marine areas through atmospheric transport [10]. The widespread use of OPEs, along with their substantial release to marine environment and resulting environmental and ecological effects, has prompted extensive scientific concern [30]. However, systematic research on OPEs in the marine environment is still insufficient, especially lacking data from open sea and deep-sea regions. Their environmental behavior, migration and transformation mechanisms, and multi-media distribution patterns require further clarification. Therefore, this review systematically summarizes the latest research advances on the sources, migration pathways, distribution characteristics, and influencing factors of OPEs in marine systems. It aims to identify the key knowledge gaps in current research and propose future research directions, with the goal of providing important scientific basis for related mechanism research, environmental policy formulation, and risk assessment.
Table 1. Structures, names, abbreviations, and physicochemical properties of commonly studied OPEs.
Table 1. Structures, names, abbreviations, and physicochemical properties of commonly studied OPEs.
CategoryCompoundAbbr.Molecular FormulaMolecular MasslgKowSolubility
(mg/L, 25 °C)		Structural Formulae
Group A: alkyl-Trimethyl phosphateTMPC3H9O4P140.07 −0.7 3.00 × 105Jmse 13 02162 i001
Triethyl phosphateTEPC6H15O4P182.16 0.8 5 × 105Jmse 13 02162 i002
Tri-n-butyl phosphateTnBPC12H27O4P266.32 4.0 280Jmse 13 02162 i003
Tri-iso-butyl phosphateTiBPC12H27O4P266.31 3.6 3.72Jmse 13 02162 i004
Tripropyl phosphateTPPC9H21O4P224.23 2.4 6450Jmse 13 02162 i005
Tris(2-ethylhexyl)phosphateTEHPC24H51O4P434.64 9.4 2Jmse 13 02162 i006
Tris(2-butoxyethyl)phosphateTBOEPC18H39O7P398.48 3.8 1100Jmse 13 02162 i007
Group B: aryl-Triphenyl phosphateTPhPC18H15O4P326.29 4.6 1.9Jmse 13 02162 i008
2-Ethylhexyl diphenyl phosphateEHDPPC20H27O4P362.40 6.3 1.9Jmse 13 02162 i009
Tri-m-Tolyl PhosphateTmCPC21H21O4P368.36 5.1 1.20 × 10−2Jmse 13 02162 i010
Tri-o-Tolyl PhosphateToCPC21H21O4P368.36 5.1 0.36Jmse 13 02162 i011
Cresyl diphenyl phosphateCDPC19H17O4P340.31 4.5 0.24Jmse 13 02162 i012
Bisphenol-A bis(diphenyl phosphate)BDPC39H34O8P2692.63 7.4 0.42Jmse 13 02162 i013
Group C: chlorinated-Tris (2-chloroethyl) phosphateTCEPC6H12Cl3O4P285.48 1.4 7000Jmse 13 02162 i014
Tris(1-chloro-2-propyl)phosphateTCPPC9H18Cl3O4P327.56 2.6 1200Jmse 13 02162 i015
Tris(1,3-dichloro-2-propyl) phosphateTDCPPC9H15Cl6O4P430.90 3.8 1.5Jmse 13 02162 i016
Note—abbr.: abbreviation; lgKow: octanol–water partition coefficient (adapted from [8,30]).

2. Sources, Transport, and Degradation of OPEs in Marine Environments

2.1. Source of OPEs in Marine Environments

The ocean serves as a significant reservoir for organic pollutants. Consequently, it plays a crucial role in the environmental transport, fate, and sinks of OPEs at both regional and global scales [31]. Marine OPEs primarily originate from two major sources: terrestrial inputs [18,24,25,26,27,28,29] and in situ releases within the marine environment [32,33]. Among these, terrestrial sources constitute the predominant contributor of OPEs in marine systems [18]. For instance, OPEs are widely incorporated as flame retardants and plasticizers into industrial products such as plastics, furniture, electronic devices, and construction materials [8,18,34]. OPEs may be released into the environment through wastewater [35,36], exhaust gases [37], and solid waste during manufacturing, usage, and disposal processes [14,38,39] of these materials. Furthermore, the use of consumer products, including textiles, coatings, adhesives, and personal care products, also leads to environmental emission of OPEs. Additionally, tire wear particles generated during transportation, as well as fuels and lubricants, contain OPEs that can indirectly enter marine waters via surface runoff and atmospheric deposition [40,41]. Effluents from municipal wastewater treatment plants, along with pesticide residues and plastic film remnants derived from agricultural activities, may ultimately reach the ocean through fluvial transport and soil leaching [3,42].
Internal sources within the marine environment also contribute to OPEs release [32,33]. Marine plastics and microplastics may release OPEs through physical abrasion, photodegradation, and biodegradation. OPEs originating from ship coatings, hydraulic fluids, and offshore platform equipment enter seawater via corrosion and leakage. Those accumulated in seabed sediments can be resuspended and released due to disturbances such as ocean currents, dredging activities, or bioturbation [26,43]. Moreover, the uptake and excretion of OPEs by marine organisms facilitate their recycling within the water column [44]. In summary, the sources of OPEs exhibit both the extensive nature of terrestrial inputs and the complexity of marine endogenous releases, collectively forming their multi-source characteristics in the marine environment.

2.2. Transport of OPEs in Marine Environments

The transport of OPEs in the marine environment is a complex, multi-medium, multi-pathway process (Figure 1). Key pathways include riverine transport, atmospheric transport, sea–air exchange and oceanic circulation, which collectively determine the spatial distribution and ultimate fate of OPEs in the global ocean.
(1)
Riverine transport serves as an important pathway for the transmission of OPEs in coastal waters [30]. Terrestrial OPEs enter rivers through wastewater discharge, surface runoff, and industrial effluents, subsequently entering coastal environments via estuaries. During this process, estuaries often function as “filters”, where some OPEs undergo migration and transformation due to particle adsorption or sedimentation. The remaining portion disperses along the coast with freshwater plumes and undergoes long-distance migration driven by coastal currents and ocean currents, ultimately being transported to the open ocean [26,45,46,47]. For instance, the Mackenzie River in Canada discharges into the Arctic, where the concentration of OPEs in nearshore waters and sediments increases. This observation supports the hypothesis that riverine transport and discharge are key sources of OPEs in the Canadian Arctic [43]. Based on the annual runoff volume of Germany’s Elbe River, approximately 50 tons of OPEs enter the North Sea each year [48]. Based on annual river discharge estimates, approximately 113 tons of TPPO enter the Bohai Sea annually, with roughly 7 tons, 1.5 tons, and 3.4 tons of OPEs entering the Liaodong Bay, Laizhou Bay, and the Bohai Sea, respectively [24].
(2)
Atmospheric transport also provides a significant pathway for OPEs entering the marine environment. Long-range atmospheric transport (LRAT) can carry OPEs originating from industrial zones and urban clusters to remote marine areas [49]. The atmosphere serves as a medium for the long-distance transmission of these pollutants. Atmospheric OPEs primarily exist in gaseous and particulate-bound forms, entering surface seawater through dry deposition (e.g., aerosol deposition) and wet deposition (e.g., precipitation processes like rain and snow) [50,51]. Substantial data supports the existence of this phenomenon. For instance, cruise data from East Asia to the Arctic indicate that the concentration of OPEs in aerosol samples collected from the Sea of Japan can reach as high as 2900 pg/m3 [49]. The production and utilization of OPEs by Asian countries significantly influence the atmospheric concentration of these compounds in this region. Additionally, research conducted in the North Sea demonstrates that OPEs emitted from urbanized and industrialized areas in Western Europe are transported to the North Sea via LRAT [52]. However, it is worth noting that a consensus on the duration for OPEs can stably exist in the atmosphere has not yet to be reached [30]. Earlier studies suggested OPEs as a potentially “environmentally friendly” alternative for PBDEs due to their supposed low persistence in the environment and presumed lack of LRAT [43,53]. In contrast, recent research indicates that OPEs can facilitate long-distance transport over extended time scales [50,54]. Data comparing different compound structures reveal that long-distance transmission is primarily associated with chlorinated OPEs [54]. This finding implies that the transport of OPEs is closely related to variations in their physical and chemical properties, as well as their stability and degradation characteristics in both air and water [30].
(3)
Air–water exchange is the key interfacial process governing the transfer of OPEs from the atmosphere into seawater. This process involves multiple mechanisms including gas diffusion, dry deposition, and wet deposition [55]. Gaseous OPEs can enter seawater via gas–liquid molecular diffusion, while OPEs adsorbed onto particulate matter enter the ocean directly through deposition. Dry deposition fluxes of OPEs ranging from 4 to 140 ng/m2/d have been observed in tropical and subtropical regions of the Atlantic, Pacific, and Indian Oceans [54]. Particulate OPEs observed in the atmosphere across different ocean regions indicate that atmospheric dry deposition [37] can remove particulate OPEs from the atmosphere. Rainfall and snowfall effectively remove OPEs from the atmosphere and rapidly transport them to the ocean surface [47,56,57,58]. Highly water-soluble OPEs are also readily removed by wet deposition. High concentrations of OPEs have been observed in some terrestrial rainwater samples [52,59], underscoring the significance of wet deposition. In high-latitude oceans, snowfall effectively removes atmospheric OPEs over the Arctic, Southern Ocean, and Antarctica. High OPEs concentrations observed in snow samples from Arctic and Antarctic expeditions underscore the significance of snowfall deposition [56,59]. Collectively, these processes facilitate the transfer of OPEs from the atmosphere to marine environments and significantly influence the subsequent accumulation of pollutants and their transfer through food web within marine ecosystems [57].
(4)
Seawater can also serve as a transport medium for organic pollutants, facilitating the long-distance movement of OPEs. Compared with the open seas, coastal seas will have higher levels of OPEs based on the receiving large inputs of water-bound OPEs [15,60]. Once in the ocean, OPEs can be subject to long-range transport via oceanic circulation [30]. For example, the transport of OPEs in seawater, driven by seawater dynamics, appears to be confined to the Canadian Arctic [43] and the North Atlantic [26]. Additionally, OPEs enter aquatic environments through atmospheric deposition and riverine input, subsequently undergoing a series of processes before being sequestered in deep-sea sediments. Dissolved organic pollutants tend to adsorb onto particles or plankton, and through the sinking of these particles and the vertical migration of zooplankton, they are removed from surface waters and transported to the deep sea for storage in seabed sediments [61,62]. Consequently, OPEs accumulate in seabed sediments, forming repositories in various marine regions. A significant quantity of OPEs has been detected in the sediments of the Yellow Sea, Bohai Sea, and other locations in China [15,63,64]. Additionally, the five-year contribution of OPEs in the sediments of the Canadian Arctic continental shelf has reached 4.1 × 103 tons [43]. These findings underscore the importance of monitoring and assessing the OPEs pollution in marine fauna, particularly in remote sea areas, deep-sea basins, and trenches.
Furthermore, global climate change can also significantly influence OPEs transport behavior. Melting polar and mid-latitude glaciers release OPEs previously frozen in ice and snow back into seawater and surrounding environments. This process not only alters regional pollutant composition but may also enhance sea–air exchange fluxes by modifying fugacity gradients [65]. For example, along the eastern Greenland coast, freshwater influx from melting ice results in local seawater OPEs concentrations that are 2–5 times higher than those in the open waters of the Fram Strait [30,56], demonstrating the potential amplifying effect of climate feedback on pollutant transport [63,66].
In summary, the migration pathways of OPEs in the marine environment are complex and their impact is widespread. Therefore, there is an urgent need to strengthen systematic research in remote areas such as the open ocean, deep sea, and polar regions.

2.3. The Environmental Degradation of OPEs in Marine Environments

The degradation of OPEs in marine environments is a critical component of their fate and behavior, primarily involving three pathways: hydrolysis, photochemical degradation, and biodegradation. These processes influence the persistence, mobility, and ecological risks of OPEs.
(1)
The hydrolysis of OPEs fundamentally involves a nucleophilic substitution process [67]. There are two primary mechanisms of hydrolysis: one in which hydroxide ions (OH−) or water (H2O) attacks the phosphorus (P) atom, and another in which water (H2O) attacks the carbon (C) atom [68,69]. The cleavage of ester bonds during the hydrolysis process is primarily influenced by the pH value of the reaction solution and the structure of the side chain [11,45]. With increasingly basic pH the order of OPEs stability was group A (with alkyl moieties) > group B (chlorinated alkyl) > group C (aryl) [11]. For group A with alkyl moieties, an increase in the number of carbon atoms in the chain corresponds to a decrease in the rate of hydrolysis [54,56].
(2)
Photochemical degradation represents a significant transformation pathway for organic pollutants in aquatic environments [10]. The photochemical degradation of OPEs in water can be categorized into direct and indirect photolysis [67]. Direct photolysis occurs when OPEs absorb light energy directly [70], whereas indirect photolysis involves the absorption of light energy by other chemical substances, such as natural organic matter, which subsequently transfer energy to transient reactive species (e.g., OH and singlet oxygen) that induce the degradation of OPEs [71]. Chlorinated and alkyl OPEs are particularly resistant to direct photolytic degradation, primarily due to their lack of absorption bands across the ultraviolet-visible spectrum, with only aryl OPEs exhibiting absorption bands within the range of 220–400 nm [72]. In real marine environments, indirect photolysis often plays a more significant role due to the influence of water transparency and dissolved substances.
(3)
Biodegradation is one of the key pathways for OPEs transformation in the environment, particularly in water bodies rich in microorganisms (such as rivers, estuaries, and sewage discharge areas) [73]. This degradation typically relies on microbial enzyme systems for the stepwise hydrolysis of phosphate bonds, involving sequential catalysis by phosphotriesterases, phosphodiesterases, and phosphomonoesterases [30,74]. To date, the only identified phosphotriesterase that facilitates the biodegradation of TCEP and TDCIPP is a haloalkylphosphorus hydrolase (HAD) [75]. And some recent studies have indicated that the presence of the nutrient element phosphorus (P) in water may influence the stability of these compounds [75,76,77].
In summary, the degradation of OPEs in marine environments is a complex, multi-pathway process whose efficiency is jointly influenced by compound structure and environmental conditions (e.g., pH, light exposure, microbial communities, and nutrient levels). The relative contributions of different degradation pathways in marine environments and the coupled effects of multiple processes require further investigation. Particularly, little is known about degradation behavior in extreme environments such as the deep sea and polar regions.
Jmse 13 02162 g001
Figure 1. Major sources and pathways of OPEs in coastal and open oceans (adapted from [30,78]).

3. Distribution of OPEs in Different Marine Environments

OPEs are generally semi-volatile compounds with relatively high solubility. They have been extensively detected in primary environmental media [6]: the atmosphere, seawater, and marine sediments. The concentrations of OPE pollutants in these major media across various sea areas worldwide are as follows:

3.1. OPEs in Atmosphere

OPEs are semi-volatile organic compounds (SVOCs) that can be transported regionally and globally through the atmosphere. OPEs have been detected in major oceans, including the Pacific, Indian, Atlantic, and Arctic [37,49]. Reported concentrations of OPEs in global marine aerosol samples range from picograms per cubic meter (pg m−3) to several nanograms per cubic meter (ng m−3), with TCEP and tris(1-chloro-2-propyl) phosphate (TCIPP) being the predominant compounds [49]. Overall, airborne concentrations of OPEs are elevated in nearshore areas, particularly those adjacent to urban and industrial regions, followed by mid-latitude open oceans and the Arctic, with the Southern Ocean exhibiting the lowest concentrations [37,79]. The global prevalence of OPEs sharply contrasts with predictions from previous LRAT models. This discrepancy suggests that our understanding of the physicochemical properties of OPEs remains significantly inadequate [30,80].
The concentration of OPEs in the Arctic Ocean at high latitudes is as low as 29–180 pg m−3; however, near Russia and the United States, concentrations are significantly higher, reaching approximately 600 pg m−3 [56,57]. In contrast, concentrations in the North Sea, Mediterranean Sea, and Black Sea at mid-latitudes can reach ng m−3 [65]. OPEs concentrations in subtropical and tropical oceans are comparable, while those in the northern and southern oceans are of similar magnitudes [54]. Notably, OPEs concentrations along the coast of the Northern Hemisphere exceed those of the Southern Hemisphere. For instance, the concentration from the Sea of Japan to the North Pole ranges from 230 to 2900 pg m−3, which is higher than the concentrations observed in aerosol samples from East Asia, the Indian Ocean, and the Antarctic, which range from 120 to 1700 pg m−3 [49]. Generally, OPEs concentrations detected in aerosol samples from coastal stations are greater than those found in open ocean samples [54,77] (Figure 2). For example, the concentration of OPEs in aerosol samples collected along the Yellow and Bohai Seas in 2020 ranged from 426 to 3933 pg m−3 [81], whereas concentrations in marine aerosol samples from the Yellow and Bohai Seas and Beihuangcheng Island during 2015–2016 were notably lower, ranging from 44 to 520 pg m−3 [37].
Figure 2. OPEs in the ocean air (adapted from [10]).
The distribution of OPEs in the atmosphere exhibits several notable characteristics. Chlorinated OPEs are the most abundant compounds, with TCEP and TCPP concentrations surpassing those of TDCP [74,82]. Additionally, the levels of TPPO and TiBP species are relatively high in urban areas along the Bohai Sea and the Mediterranean Sea, primarily due to local pollution sources associated with OPEs [40]. One reason for the elevated presence of chlorinated OPEs in the atmosphere compared to other OPE congeners is their longer photodegradation time [18]. The proportion of OPE homologues in the atmosphere of various bordering countries correlates with the usage of these compounds within those nations. For instance, European countries have prohibited the use of TCEP as an additive, resulting in the detection of higher levels of TCPP than TCEP in nearby marine environments [65]. Conversely, TCEP remains in use in Asian markets, where its concentration in coastal atmospheres exceeds that of TCPP [49,82]. Furthermore, the increasing TCPP/TCEP ratio in Asian waters indicates a rise in TCPP usage within these countries [74].

3.2. OPEs in Seawater

Currently, with the increasing emphasis on marine environmental research, surveys of OPEs in seawater samples have been conducted to varying extents in different worldwide regions [46,83] (Figure 3 and Table 2). The findings indicate that OPEs have been detected in seawater samples from major oceanic areas, highlighting the long-distance migration potential of these compounds [30,48]. The concentrations of OPEs identified in various seawater environments exhibit significant variability, ranging from a few to several hundred ng L−1 [25,84]. A recent review article synthesized the distribution of OPEs across key sea areas globally [30,85]. Notably, by analyzing coastal survey data from seven European countries [86], it was determined that OPEs were present in all ocean surface water samples collected from these regions. The total concentration range was found to be between 0.43 and 867 ng L−1 [48,87], with elevated levels primarily observed near urban areas in the UK and Portugal.
Figure 3. Concentrations of OPEs (ng/L) in seawater. (a) Concentrations of different OPEs compounds. (b) The total amount of OPEs. Data sources are provided in Table 2.
Seawater survey data from the western Pacific Ocean indicate that the concentration range of OPEs is between 3.02 and 48.4 ng L−1 [88]. Similarly, studies of water samples from various coastal areas in China and Japan have demonstrated the widespread distribution of these compounds in these maritime regions. Samples collected in 2016 from China’s Bohai Sea, Yellow Sea, and East China Sea revealed concentration ranges of 19.7 to 100 ng L−1, 9.26 to 86.8 ng L−1, and 8.81 to 55.7 ng L−1, respectively [10,12,34,84,89]. Additionally, the distribution range reported by [90] for the South China Sea is 1 to 147 ng L−1, while the concentration range for Tokyo Bay in the Sea of Japan is 107 to 284 ng L−1.
Chlorinated OPEs have also been detected in higher latitude seas [30], including Arctic waters, with concentrations ranging from approximately 0.9 to 17 ng L−1 [46,91]. Similarly, the concentrations of ∑8OPEs in the North Atlantic and Arctic were found to range from 0.35 to 8.4 ng L−1 [56]. In studies conducted near the northwest Pacific and Arctic waters, the distribution of OPEs concentrations in samples varied from 8.47 to 143.45 ng L−1 [57]; however, only one sample reached 143.45 ng L−1, while the values of the remaining samples were all below 50 ng L−1. Unfortunately, the article did not provide an explanation for this unusually high value.
It is important to note that the primary compounds detected in seawater are TCPP, TCEP, and TDCPP. This distribution characteristic can be attributed to the extensive use of these three chlorinated OPEs. Furthermore, their stable branched chain structures and high resistance to degradation also contribute to their persistence in seawater [85].
The factors influencing the distribution characteristics of OPEs in seawater are complex. Comparative research indicates that the concentration of OPEs in seawater samples varies across different sampling times, depths, and locations. Research conducted in China’s Bohai Sea, South China Sea, and various other marine regions has demonstrated that the concentration of OPEs in seawater exhibits a gradient distribution from estuarine to coastal and pelagic zones [10]. As the distance from shore increases, the concentration of OPEs gradually diminishes [24,92]. These findings highlight the influence of sampling location on the distribution patterns of OPEs.
A comparison of data from the Yellow Sea in 2015 and 2019 revealed notable changes in OPE concentrations across different years, with varying patterns observed among different compounds [12]. For instance, TCPP exhibited a significant increase, while TCEP showed no substantial change. The rising use of TCPP as a substitute for TCEP may explain these trends. More importantly, a comparison of the total amounts of chlorinated OPEs in the Yellow Sea, Bohai Sea, and other seawater bodies over the past five years indicates an exponential increase [60]. These data illustrate that chlorinated OPEs can persist stably in the ocean for extended periods. Conversely, they also indicate an annual increase in emissions from adjacent coastal areas.
Table 2. Occurrence of OPEs in seawater (average value, ng/L).
Table 2. Occurrence of OPEs in seawater (average value, ng/L).
Country/RegionLocationSampling
Year		TCEPTCPP/TCIPPTDCPP/TDCPTBEPTnBPTPhPTiBPTEHPTEPTPPOEHDPPΣOPEsRef.
ChinaMariculture area in the Beibu Gulf201513.1 (5.21–82.1)28.2 (13.9–92.5)0.49 (0.19–1.68)na1.53 (0.69–4.8)3.99 (1.28–7.76)2.74 (0.63–5.68)0.14 (0.03–0.36)nanana68.7 (32.9–227)[93]
ChinaMariculture area of Hebei201731.97 (4.17–74.76)34.39 (14.49–69.68)1.04 (0.52–3.01)ndna0.65 (0.11–3.22)nanananana74.5 (40.40–154.05)[34]
NorwayNy-Ålesund region20165.58 (nd–60.84)2.45 (nd–15.03)0.34 (nd–0.99)ndna0.25 (nd–0.89)na0.79 (nd~11.37)nanana13.4 (0.66–61.64)[34]
USASan Francisco Bay20137.4–30046–290014–45024–10007.8–4341–360nand–11nd–3.2nand–2.3170–5100[83]
GermanyNorth Sea2010na3–28nand–6nana0.5–5na0.7–7nd–12na5–50[48]
OceanPacific Northwest to Arctic Ocean20189.9 (1.10–86.19)3.81 (0.76–20.87)1.25 (nd–4.53)0.32 (nd–7.70)0.02 (nd–0.33)0.88 (0.26–2.64)4.07 (2.37–6.00)0.37 (0.0004–1.55)nanana8.47–143.45[57]
OceanDeep-water system of the North Atlantic2014–20150.08 (nd–0.39)0.04 (nd–0.05)0.004 (0.001–0.007)na0.01 (nd–0.06)0.60 (nd~1.2)na0.61(nd–1.5)nana0.21(0.06–0.33)0.0063–0.44[46]
CanadaBarrow Strait2014–20150.001 (0.0001–0.002)0.003 (0.0001–0.006)0.00006 (nd–0.0001)na0.0004 (0.0002–0.0006)0.0006 (0.0004–0.0008)na0.00002 (nd–0.00005)nana0.0003 (nd–0.0006)-[46]
OceanNorth Atlantic and Arctic Oceans20140.70 (nd–2.40)1.84 (0.28–5.77)0.007 (nd–0.04)na0.12 (nd–0.41)nd0.26 (0.04–0.64)0.006 (nd–0.07)nanana2.94 (0.34–8.59)[1]
ChinaBohai Sea20169.56 (6.06–19.8)11.4 (3.97–26.7)2.03 (nd–5.16)nand–37.20.15 (nd–3.28)6.96 (1.97–27.4)nana5.16 (1.86–26.6)na39.1 (19.7–100)[15]
Yellow Sea5.84 (1.24–16.9)13.1 (5.17–35.6)2.07 (nd–8.13)nand–26.50.14 (nd–0.76)nd–9.64nana6.93 (1.18–43.5)na30.7 (9.26–86.8)
East Sea1.93 (0.59–12.4)9.63 (5.61–29.6)0.51 (nd–4.92)nand0.11 (nd–1.95)ndnana1.63 (nd–19.1)na15.1 (8.81–55.7)
ChinaYellow Sea2022nanand34.51 (nd–334.88)2.61 (nd–9.36)ndnanananana22.94 (nd–497.40)[89]
East Seanana7.73(nd–9.84)30.15 (nd–124.82)2.12 (nd–22.74)ndnanananana8.11 (nd–126.49)
ChinaYellow Sea and East Sea2023nanananananananananana66.73 (12.72–202.60)[12]
ChinaPearl River2022169.4 ± 157.1 (39.3–414.9)105.9 ± 78.6 (37.8–260.9)7.4 ± 3.9 (1.2–11.6)na50.3 ± 57.7 (13.6–179.0)1.8 ± 1.5 (0.7–4.2)nanananana361.8 ± 283.3 (117.5–854.8)[10]
Northern part of the South Sea2.0 ± 1.9 (0.2–5.4)2.7 ± 3.2 (0.3–10.3)0.06 ± 0.04 (0.002–0.1)nana0.09 ± 0.07 (0.01–0.2)na0.08 ± 0.05 (0.01–0.2) 0.05 ± 0.02 (0.01–0.09) 6.7 ± 5.2 (1.3–17.6)
Western part of the South Sea4.2 ± 4.7 (0.8–22.6)2.4 ± 1.7 (0.4–6.2)0.06 ± 0.08 (nd–0.3)na0.5 ± 0.5 (0.1–2.2)0.06 ± 0.03 (0.01–0.10)nanananana7.6 ± 5.5 (2.3–24.4)
na: not analyzed; nd: not detected.
In addition, Zhang Lutao [12] compared the distribution of OPEs in seawater samples collected at various depths in the Yellow Sea. The results highlighted differences in the concentration distribution of OPEs among surface seawater, 20 m deep seawater, and bottom seawater samples, indicating the influence of sampling depth.
To fully understand the distribution characteristics of OPEs in seawater, it is essential to conduct comparative studies under varying conditions. Such studies will provide a comprehensive and accurate understanding of the distribution, physical and chemical properties, as well as the migration and transformation mechanisms of these compounds.

3.3. OPEs in Marine Sediment

Currently, there is a limited amount of research data on marine sediments, primarily focused on marginal seas near inland areas, with significantly less data available from the open ocean. Published data indicate that the total concentration of OPEs varies from pg g−1 dw to ng g−1 dw [10,30] and the following distribution characteristics are present (Figure 4 and Table 3). Furthermore, a trend of increasing OPEs concentrations in marine sediments is observed, progressing from polar regions to estuaries. For instance, the concentrations in the Bering Strait and the Arctic Ocean are reported to be 0.29–2.01 ng/L and 0.32–4.66 ng/L [63], respectively. The distribution ranges of Yellow Sea and Bohai Sea are 0.08–1.86 ng/L and 0.21–4.55 ng/L [94]. In contrast, the Beibu Gulf marine aquaculture area exhibits a distribution range of 4.35–22.1 ng/L. Additionally, the distribution ranges in the Taiwan Strait and Qiongzhou Strait are 5.26–34.23 ng/L and 0.99–36.2 ng/L [95,96]. And the available research data reveal several consistent characteristics.
Figure 4. Concentrations of OPEs (ng/g) in marine sediments. (a) Concentrations of different OPEs compounds. (b) The total amount of OPEs. Data sources are provided in Table 3.
(1)
Chlorinated OPEs, particularly TCEP and TCIPP, are the most abundant components in sediments located closer to the continent [10,30]. This prevalence is attributed to their widespread use and low degradation rates. Furthermore, due to their relatively high hydrophobicity, TEHP (lgKow 9.5) and TCrP (lgKow 5.1) were also identified as dominant components of OPEs in the sediments [94,95].
(2)
The levels of OPEs in sediments tend to be higher in areas closer to land and more densely populated regions [30]. In contrast, the concentrations of OPEs in open-ocean sediment are significantly lower than those found in marginal-sea sediments [97]. For instance, in the Bohai Sea, decreasing levels of ΣOPEs have been reported in sediment as the distance from shore increases: Laizhou Bay (6.6–100 ng g−1 d.w.) > Bohai Bay (1.7–29 ng g−1 d.w.) > Bohai Sea (0.20–4.6 ng g−1 d.w.) [98,99]. This pattern indicates that the primary sources of OPEs in marine sediments are likely terrestrial contributions. Conversely, the concentrations of OPEs in open-sea sediments are significantly lower, such as those from the North Pacific to the Arctic Ocean, range from 0.2 to 4.7 ng g−1 d.w [63].
(3)
Differences in the content of OPEs in sediments can be attributed to varying depths. Investigating the concentration of organic pollutants in sedimentary columns at different depths is essential for understanding the pollution history across different time periods [100]. Several factors are currently recognized as influencing the vertical distribution of OPEs, including variations in regional production, usage, and emission intensity within the current year [99,101]. Additionally, differences in the adsorption and distribution behavior of various compounds in water and sediments arise from their distinct physical and chemical properties, as well as the transformation and degradation of chemical compounds during post-deposition. Nonetheless, there remains a significant gap in systematic research within this field, leading to an insufficiently comprehensive summary of existing patterns. Therefore, further investigation is warranted.
Table 3. Occurrence of OPEs in ocean surface sediment (average value, ng/g dw).
Table 3. Occurrence of OPEs in ocean surface sediment (average value, ng/g dw).
Country/
Region		LocationSampling
Year		Sampling
Number		TCEPTCIPPTDCIPPTEPTnPPTnBPTiBPTPePTEHPTBOEPTPHPTCrPEHDPPΣOPFRs (Range)Ref.
ChinaBeibu Gulf2015120.8802.350.090 0.2101.33 0.3600.1600.6900.040 7.56 (4.35–22.1)[63]
ChinaTaiwan Strait2016321.101.310.124ndnd7.92 1.820.3430.5490.042 12.8 (5.26–34.2)[102]
ChinaCoast of Hainan Island
(Qiongzhou Strait)		2015203.272.690.8700.6400.3800.660 1.15nd0.7604.670.31015.9 (0.990–36.2)[30]
Coast of Hainan Island (Near-shore)141.623.280.5500.7900.2301.04 0.850nd0.5107.950.48016.4 (nd–60.0)
coast of Hainan Island
(Off-shore)		91.602.720.6401.020.8801.34 0.870nd0.5806.050.41015.1 (2.40–28.4)
ChinaLaizhou Bay, Bohai Sea2017–2018157.4016.320.115.09.7033.811.026.032.168.214.213.618.4304.2 ± 16.2[98]
ChinaBohai Bay, Bohai Sea2014–2017120.6371.480.0550.346nd0.184 0.2320.1890.0820.0910.0683.79 (1.66–28.7)[30]
ChinaBohai
Yellow Seas		201044
43		0.202
0.111		0.113
0.076		0.018
0.011		 0.024
0.010		0.085
0.016		0.002
0.002		0.375
0.080		 0.053
0.036		 1.14 (0.205–4.55)
0.411 (0.083–1.86)		[103]
ChinaPearl River Estuary20101013.04.300.3400.4000.0400.600 4.203.50nd-16.00.3000.62034.0 (12.0–66.0)[104]
ChinaYellow Sea202266 9.02 1.16 5.28 (nd–47.85)[89]
East Sea 5.31 1.60 6.10 (nd–66.50)
JapanMaizuru Bay2009133.00 13.0nd 3.00 7.002.003.00 <0.500–56.0[48]
KoreaKorean coast2016502.1641.51.781.70 1.73 5.957.631.831.33/3.94/7.921.7871.0 (2.18–347)[40]
NetherlandsWestern Scheldt estuary200830.5003.600.600 6.90 1.509.500.6000.4000.300-[105]
USASan Francisco Bay201410nd0.5400.960nd-0.570 8.200.8101.903.400.28023.0[47]
Northwest
Mediterranean Sea		Gulf of Lion201812 7.7825.05 2.550.598.85 7.36 0.92 1.9754 (4–227)[43]
ArcticNy-Ålesund20179<0.02–2.880.01–7.41<0.02–0.73 <0.02–2.64<0.01–0.37<0.01–0.64<0.01–0.872.44 (0.01–14.94)[106]
OceanNorth Pacific to the Arctic Ocean2010300.5360.0680.014 0.0680.1620.007 0.023 0.878 (0.159–4.66)[107]
Bering Sea40.6550.0810.011 0.1140.3150.002 0.030 1.21
Bering Strait40.1090.009nd 0.0600.1700.001 0.001 0.350
Chukchi Sea120.2810.0330.006 0.0520.1260.002 0.023 0.524
Canadian Basin margin30.7170.1140.037 0.0950.1770.020 0.028 1.19
Central Arctic Ocean71.070.1360.028 0.0610.1230.015 0.028 1.46
nd: not detected; a median value; b TmCP/ToCP/TpCP.

4. Conclusions and Future Perspectives

As a result of their widespread application, OPEs have been detected at different concentration levels in marine environments globally, and OPEs’ persistence allows these compounds to exert a lasting and extensive impact on the environment and ecosystem. Consequently, it is urgent and important to investigate the distribution, migration, fate, and other related issues concerning these compounds. Based on the comprehensive analysis in this article, the following conclusions can be drawn:
(1)
OPEs in the marine environment primarily originate from terrestrial sources and in situ releases. The transplantation of OPEs are complex, including river input, LRAT, air–water exchange, and dry and wet deposition. Additionally, OPEs can be remobilized by snow and ice melting, and they can also be transported in ocean via currents and biological vectors (e.g., food chains). Furthermore, OPEs are found to be degraded in the environment primarily through hydrolysis, photolysis, and biodegradation.
(2)
The distributions of OPEs in marine environments have been intensively investigated. OPEs have been detected in the air of major oceans, with concentrations in global marine aerosol samples ranging from pg m−3 to several ng m−3. Notably, chlorinated OPEs are the most abundant compounds identified. Furthermore, OPEs have been found in seawater samples from various regions, confirming their long-distance migration experience. The concentrations of OPEs identified in seawater exhibit significant variability, ranging from a few to several hundred ng L−1. The primary compounds detected in seawater are TCPP, TCEP, and TDCPP. The data published indicate that the total concentration of OPEs in marine sediments varies from pg g−1 dw to ng g−1 dw. Furthermore, an increasing trend of OPE concentrations in marine sediments is observed from polar regions to estuaries. Chlorinated OPEs, particularly TCEP and TCIPP, are identified as the most abundant components in sediments located closer to the continent. The concentration levels of OPEs in sediments tend to be higher in areas closer to land and more densely populated regions.
Future studies might be conducted in the following directions for the purposes of ecological conservation and to deepen our understanding of marine biogeochemical cycles:
(1)
Systematic investigations of OPEs in different marine environments: It is essential to carry out systematic comparative studies, including analyses of air, seawater, and sediments within the same area, to advance understanding of the migration and transformation processes of OPEs in marine systems. Comparisons should also be made to understand the influences of environmental factors on the distributions of OPEs. To clarify the mechanisms of their environmental distribution differences, the structural details of OPEs should be specified.
(2)
Development of predictive models: Establishing a machine learning framework, which integrates key physicochemical parameters (e.g., lgKow, pKa, and hydrolysis rate constants), oceanic environmental variables (e.g., ocean current dynamics and meteorological conditions), and annual organic phosphate emissions (e.g., flux, geographic distribution, and chemical composition), could aid in the precise prediction of the fate of OPEs in marine systems.

Author Contributions

X.X.: Conceptualization, Formal analysis, Writing—Original draft, Funding acquisition. M.P.: Visualization, Writing—review and editing. Y.W.: Writing—review and editing. B.S.: Resources, Project administration. P.F.: Writing—review and editing. J.Y.: Visualization. H.L.: Conceptualization, Writing-review and editing. All authors have read and agreed to the published version of the manuscript.

Funding

This research was funded by the National Natural Science Foundation of China (Grant No. 41702159) and the Chinese Academy of Geological Sciences Basal Research Fund (Grant Nos. CSJ-2024-05 and CSJ-2023-08).

Data Availability Statement

No new data were created or analyzed in this study. Data sharing is not applicable to this article.

Conflicts of Interest

The authors declare no conflicts of interest.

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